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EXPOSURE OF MARINE BIRDS TO
ENVIRONMENTAL POLLUTANTS

SKIOIWORF^

SAR/UO0A h?mu, N, Y,

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UNITED STATES DEPARTMENT OF THE INTERIOR
FISH AND WILDLIFE SERVICE

Wildlife Research Report 9

WILDLIFE RESEARCH REPORTS

This series comprises reports of research relating to birds, mammals, and
other wildlife and their ecology, and specialized bibliographies on these, issued
for wildHfe research and management specialists. The Service distributes these
reports to official agencies, to libraries, and to researchers in fields related to
the Service's work; additional copies may usually be purchased from the
Division of Public Documents, U.S. Government Printing Office.

Library of Congress Cataloging in Publication Data

Ohlendorf, Harry M
Exposure of marine birds to environmental pollutants.

(Wildlife research report; 9)

Bibliography: p.
1. Sea birds— Physiology. 2. Pollution -Environmental aspects. 3. Pollu-
tion—Toxicology. 4. Oil spills and wildlife. I. Risebrough, Robert W., joint
author. II. Vermeer, Kees, joint author. III. Title. IV. Series.
QL698.044 598.2 '2 '4 78-5240

Use of trade names does not imply U.S. Government endorsement of commercial products.

EXPOSURE OF MARINE BIRDS TO ENVIRONMENTAL
POLLUTANTS

By Harry M. Ohlendorf
Robert W. Risebrough
Kees Vermeer

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From the collection of

International

Bird Rescue

Research Center

Cordelia, California

in association with

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0 Prelinger

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ibrary

San Francisco, California
2006

UNITED STATES DEPARTMENT OF THE INTERIOR
FISH AND WILDLIFE SERVICE
Wildlife Research Report 9
Washington, D.C. • 1978

Digitized by the Internet Archive

in 2007 with funding from

IVIicrosoft Corporation

http://www.archive.org/details/exposureofmarineOOohlerich

Contents

Page

Abstract 1

Petroleum Hydrocarbons 2

Sources of Oil in United States Waters 3

Transfer and Dissipation of Oil in the Marine Environment 4

Exposure of Marine Birds to Oil 5

Biological Effects of Oil on Marine Birds 5

Feather-oiling 6

Toxicology, physiology, and pathology 7

Reproduction 8

Behavior 9

Organochlorines 9

Exposure of Marine Birds to Organochlorines 10

Biological Effects of Organochlorines on Marine Birds 15

Toxicology, physiology, and pathology 15

Reproduction 18

Behavior 22

Heavy Metals 22

Exposure of Marine Birds to Heavy Metals 23

Biological Effects of Heavy Metals on Marine Birds 25

Toxicology, physiology, and pathology 25

Reproduction 27

Behavior 29

Plastic and Other Artifacts 29

Recommendations 30

Acknowledgments 31

References 31

Exposure of Marine Birds to
Environmental Pollutants^

by
Harry M. Ohlendorf

U.S. Fish and Wildlife Service

Patuxent Wildlife Research Center

Laurel, Maryland 20811

Robert W. Risebrough and Kees Vermeer

Canadian Wildlife Service

Department of the Environment

Ottawa, Ontario (RWR)

Delta, British Columbia (KV)

Abstract

It is unlikely that any marine birds remain uncontaminated by the synthetic
organochlorine compounds that have become ubiquitous pollutants. Marine
birds also are increasingly exposed to petroleum compounds as a result of the
exploitation of undersea petroleum deposits, increased tanker traffic, and ex-
pansion of coastal petrochemical industries.

Lethal and reproductive effects of organochlorines on marine birds have
been most pronounced in coastal areas receiving effluents discharged by
manufacturing plants. For example, particularly severe DDT contamination in
southern California and elevated levels of dieldrin and related chemicals in the
Netherlands have killed local marine birds or inhibited their reproduction.
Eggshell thinning, apparently resulting from exposure to DDE, is widespread
among estuarine species, and eggshells of peregrine falcons {Falco peregrinus)
have become thinner in all areas of the species' range thus far studied. In
more contaminated coastal areas, reproductive success of the peregrine falcon
is low. Adverse effects of organochlorines on the reproduction of other species
also have been found.

The oihng of feathers and the associated mortality of marine birds are not
the only adverse effects of oil pollution; ingestion of oil may cause death by
dehydration by interfering with ion transport and water balance in the gut.
Surfactants used to disperse oil spills also have serious consequences for ma-
rine birds. Dissolved oil fractions may kill or poison the biota the birds feed
on. The physiological effects of the incorporation of more persistent com-
pounds into marine food webs are unknown.

Contamination of marine birds by most metals and certain trace elements
has not been documented, although elevated mercury levels have been ob-
served in birds of certain estuarine and local marine environments. The signifi-
cance of elevated mercury levels and small plastic particles found in the
stomachs and pellets of marine birds is not yet known.

' A summary of this paper was presented at the 13-15 May 1975 international symposium on "Conser-
vation of Marine Birds of Northern North America," in Seattle, Washington. The paper was written
in 1975-76, and certain portions have been amended or updated as references were pubHshed. Proceed-
ings of the symposium are being prepared for publication, but this paper on environmental pollutants
is being published separately because of its great length and the delay in pubhcation of the entire
Proceedings.

Marine birds are exposed to several types of
environmental pollutants: petroleum hydro-
carbons, organochlorines, heavy metals, and
others. Because few data are available for
northern North America, potential problems
for marine birds there must be judged from
observations in other geographical areas.

Certain marine birds may serve as indi-
cators of environmental pollution on a global
scale because (1) they usually can be identified
even in an advanced state of decomposition,
after a long period of submergence in sea-
water, or when thickly covered with oil; (2)
they are geographically widespread, often are
very numerous, and feed on a wide range of
marine organisms; and (3) most species nest
colonially and lay large, distinctively marked
eggs that are often easily collected and consti-
tute distinct units for comparison between
species (Vermeer and Reynolds 1970; Prestt
1971).

Eggs serve as particularly useful sample
units for analysis of organochlorines and
certain heavy metals, particularly mercury,
because they do not decompose rapidly and
are easily handled. Some marine bird species
lay additional clutches if the first is removed;
therefore, eggs may be taken without severe
adverse effects on populations. This charac-
teristic is of particular importance because
studies are sometimes not begun until it is ap-
parent that a population is declining (Prestt
1971).

Organochlorine concentrations in the egg
are about equal to whole body concentrations
found in the female at the time the egg was
laid (Keith and Gruchy 1972). Although some
microbes have the ability to metabolize or-
ganochlorine pesticides under certain condi-
tions (Matsumura 1974), putrefaction does
not significantly affect residue analysis for
DDT and its metabolites (Mulhern and
Reichel 1970). During incubation, however,
the developing embryo appears to metabolize
DDT to DDD and DDE (Abou-Donia and
Menzel 1968; Blus et al. 1974b). Chemical resi-
due concentrations can be adjusted for the
loss of moisture and lipids that occurs during
incubation (L. F. Stickel et al. 1973).

Eggs may not be the best tissue for mea-
surement of all metals, because certain heavy
metals apparently are not readily transferred
to them (Anderlini et al. 1972). This, however,
is not true of mercury. Under certain circum-

stances, feathers may be the best tissue to
analyze for mercury residues (Stickel 1971).
However, unlike liver and muscle tissue, mer-
cury residues in feathers tend to reflect body
burdens at the time the feathers were grow-
ing. The liver, which is a major organ of me-
tabolism, or muscle appear to be the best tis-
sues for measuring current exposure to mer-
cury (Backstrom 1969; Vermeer and Arm-
strong 1972b; Fimreite 1974). Other heavy
metals may be concentrated in other tissues.
For example, residues in the bones, kidneys,
and brain, as well as in the liver, appear to be
the best measure of exposure to lead (Long-
core et al. 1974b). The transfer of petroleum
hydrocarbons to eggs has not been reported,
but may be expected to occur.

Unless otherwise indicated, all chemical
residues in the present report are expressed
on a wet-weight basis.

Petroleum Hydrocarbons

Because much of the current information
concerning the significance of oil pollution in
the estuarine and marine environment has
been included in recent reviews (National
Academy of Sciences 1973, 1975a; Moore and
Dwyer 1974; Evans and Rice 1974; Vermeer
and Vermeer 1974a, 1974b; Farrington 1977),
we have avoided an extensive review here.
However, some of the general information,
taken in part from these reviews, is pertinent
to our subject and has been included along
with that more specifically related to birds.

Crude oil and petroleum products are com-
plex mixtures of chemicals with individual
compounds numbering in the tens of thou-
sands with wide molecular weight ranges
(Farrington 1977). No one method of analysis
is available that will provide reUable esti-
mates of the concentration of the entire range
of petroleum compounds, and there has yet to
be a complete analysis of a single crude oil.
Therefore, reports of the presence or absence
of petroleum pollution should be carefully
evaluated to be certain that the methods of
chemical analysis employed would indeed pro-
vide the information reported.

Vanadium and nickel are present in appre-
ciable quantities (> 100 ppm) as organometal-
lics indigenous to crude oil, and other trace
metals are picked up during production or
transportation of crude oil (Whisman and

Cotton 1971).

Oil pollutants have been detected in sedi-
ments, water, and organisms in areas of large
oil spills as well as from areas where no large
spills have occurred in past months or years
(Farrington 1977). These areas are near
sources of small spills and chronic inputs. No
more than an estimated 300 analyses for pe-
troleum pollutants in sediment, water, and or-
ganism samples have been reported in the lit-
erature exclusive of reports of visible sheens
on the water.

The scarcity of published measurements of
the extent and severity of oil pollution in sedi-
ments and organisms is probably related to
the difficulty of making meaningful analytical
measurements to detect petroleum pollution
(Goldberg 1972).

Sources of Oil in
United States Waters

The amount of oil entering the marine envi-
ronment from known sources has recently
been estimated on a worldwide basis
(5.3 million metric tons) as well as for the
United States (1.3 million metric tons; Na-
tional Academy of Sciences 1973, 1975a; Far-
rington 1977).

The largest amounts of oil come from
normal transport and refining operations and
are intentional discharges (Farrington 1977).
Accidents account for only 3% of the oil that
reaches marine waters of the United States
and for only 5% of the world total. Oil tdnker
operations account for 26 times as much oil as
offshore production in the United States and
24 times as much in the world total.

The oil that reaches the coastal waters from
rivers and from land operations accounts for
65% of the total (Farrington 1977). The oil
that reaches the oceans from the air, by dry
fallout and rain, is estimated to be less than
5% of the total.

The relative importance of the various
sources of oil entering the marine environ-
ment varies with location and time (Farring-
ton 1977). For example, a large well blowout
would introduce a massive amount of oil to a
given location and even if averaged over a 10-
year period would be the dominant source for
that geographical location.

The effect of the oil from the various

sources can be very different (Farrington
1977). For example, accidental spills may
have both immediate acute effects and long-
term chronic effects. Municipal or industrial
effluents, in contrast, may have no measur-
able immediate impact but may have long-
term chronic effects as the concentration of
the petroleum chemicals builds up in the
ecosystem.

Two important points relate to control of oil
pollution discharges (Farrington 1977):

(1) The largest source is the chronic drib-
bling of oil into the coastal zone by industrial
and municipal effluents, urban runoff, and
river runoff carrying oil from inland areas. A
substantial amount of oil, therefore, will be
discharged to the coastal zone regardless of
source. This amount will increase as oil con-
sumption increases unless control steps are
taken. Evidence suggests that chronic low-
level pollution could be potentially more dam-
aging to ecosystems than isolated cata-
strophic spills (Evans and Rice 1974).

(2) It may be safer for the total marine envi-
ronment to drill and produce oil in offshore
areas than to import equal quantities of oil.
Approximately 0.014% of the oil produced
offshore is discharged to the marine environ-
ment, in contrast to about 0.16% of the oil
transported by tanker. However, this does
not take into account the ecological damage
that may occur in coastal areas as a result of
the construction and maintenance of pipelines
and onshore facilities.

Mystery oil spills, those of unknown source,
account for 30% of the oil spills in U.S. waters
(National Academy of Sciences 1973, 1975a).
There are two possible ways to identify mys-
tery oil spills. The first method is to tag oil
tanker cargoes, pipeline loads, and storage
tank contents with microscopic spheres or
special chemicals. However, the size of the bu-
reaucracy necessary to ensure accurate rec-
ords renders this method impractical (Far-
rington 1977).

The second method is to make detailed
chemical analysis of the spilled oils and poten-
tial sources. The chemical characteristics are
then compared and the best match of a poten-
tial source with the spilled oil is attempted.
This technique, which is called "passive
tagging," makes use of the unique chemical
composition of each oil to distinguish one
from another and to match oils from source

and spill samples. The technique is also re-
ferred to as "fingerprinting," which is
perhaps unfortunate. Many nonscientists in
the field of oil pollution control have mistak-
enly equated "fingerprinting" in identifying
mystery oil spill sources with fingerprinting
in criminology. Although there are some simi-
larities, the identification of oils is very diffi-
cult and in its infancy as a technique (Lee et
al. 1974; Farrington 1977). However, follow-
ing an extensive investigation by the U.S.
Coast Guard, charges have been filed in the
first case that was based on chemical similari-
ties in spilled oil and a sample taken from a
suspect tanker (Anon. 1975).

Transfer and Dissipation of
Oil in the Marine Environment

A basic understanding of the various path-
ways of transfer and fate of oil has been de-
rived from laboratory studies, field studies,
and the application of knowledge of processes
in the marine environment (Farrington 1977).

Many of the processes that act on the oil re-
sult in a fractionation and selective removal of
certain components more rapidly than others
(Farrington 1977). Lower molecular weight
components of the type found in kerosene,
gasoline, and in varying concentrations in
crude oils and fuel oils will evaporate more
rapidly than the heavier molecular weight
components such as those that make up the
bulk of lubricating oils. The lower molecular
weight components also are more soluble than
the heavier components (Moore and Dwyer
1974; Farrington 1977). When oil is placed in
contact with seawater, the lower molecular
weight aromatic hydrocarbons are dissolved
or accommodated in the water to a greater
extent than are the saturated hydrocarbon
components (Boy Ian and Tripp 1971; Frank-
enfeld 1973; Boehm and Quinn 1974; Ander-
son et al. 1974a, 1974b; Lee et al. 1974;
American Petroleum Institute 1973). When a
spill occurs, however, oil may enter marine
sediments and be released essentially un-
changed months later (Blumer et al. 1970).

Extensive laboratory research has been di-
rected toward a better understanding of the
biodegradation of oil, and of the individual
compounds or classes of compounds in oil
(Davis 1967; ZoBell 1969; Ahearn and Meyers
1973; National Academy of Sciences 1973,

1975a). Several species of microorganisms,
e.g., bacteria and yeasts, will completely de-
grade certain components of oil under the
right conditions in the laboratory or in the
field (Farrington 1977).

Bacteria capable of partially degrading oil
have been isolated from several locations in
the world's oceans (Farrington 1977). How-
ever, the rates of degradation in the various
types of coastal areas are unknown. The po-
tential pathogenicity of some species of bac-
teria that might increase in number near or in
an oil spill area also is unknown and little is
known about the toxicity of the chemicals
produced by microbial degradation of oil (Na-
tional Academy of Sciences 1975a). Knowl-
edge of the biochemical pathways and prod-
ucts of the biochemical degradation of oil is
only rudimentary (Davis 1967; ZoBell 1969;
Ahearn and Meyers 1973; National Academy
of Sciences 1973, 1975a; Farrington 1977).

Oil may enter marine organisms by inges-
tion of contaminated food and may also enter
from water across membrane surfaces such as
gills (Farrington 1977).

Oil incorporation into some shellfish, lob-
sters, and fish is reversible to some extent
when the animals are placed in clean water for
a period of time. Most, but not all, of the oil
taken up from water by the animals was dis-
charged within weeks to months in clean
water (Blumer et al. 1970; Lee et al. 1972a,
1972b; Anderson 1973; National Academy of
Sciences 1973; Stegeman and Teal 1973; An-
derson et al. 1974b; Fossato 1975). However,
oysters exposed for 2 months to oil from an oil
spill did not appreciably reduce their oil
content even after 6 months in cleaner waters
(Blumer 1971; Blumer and Sass 1972). The
more toxic cyclic hydrocarbons were retained
longer than the less toxic straight chain com-
pounds (Blumer et al. 1970).

Fish tested in the laboratory partially me-
tabolized several different aromatic hydrocar-
bons of the type found in crude oils and fuel
oils (Lee et al. 1972b). Mussels, however, did
not metabolize these compounds under simi-
lar conditions, showing the undesirability of
extrapolating from one group of organisms to
another (Lee et al. 1972a). Equal caution is ad-
visable in extrapolating from results of tests
of those few compounds that have been tested
because differences in the molecular structure
can have profound effects on the rates at

which they are absorbed and metabohzed
(Farrington 1977).

Aside from these few examples, we have
found no studies of retention of petroleum hy-
drocarbons after oil spills. Neither have we
found studies of the uptake, retention, and
discharge of petroleum hydrocarbons taken in
with food. Some data suggest that food web
magnification of oil does not occur in certain
communities of marine organisms (Lee et al.
1972a, 1972b; Burns and Teal 1971, 1973; An-
derson 1973; Stegeman and Teal 1973). There
may, however, be magnification of the higher
boiling fractions of the contaminants higher
up in the food web (Burns and Teal 1971).

Chemical communication is highly im-
portant among marine organisms, for both
interspecific and intraspecific message sys-
tems. Because very low concentrations of or-
ganic stimuli are required for communication,
such processes are especially susceptible to in-
terference by pollutants at low concentrations
(Blumer 1971; Blumer et al. 1973; Jacobson
and Boylan 1973; Atema and Stein 1972,
1974).

Small quantities of crude oil (0.9 ml in
100 liters of sea water) interfere with some
specific, possibly chemosensory, behavior of
the lobster (Homarus americanus). The delay
period between noticing food and going after
it doubled when oil was added. The water-sol-
uble fraction of the oil alone (in the 50-ppb
range) did not have a noticeable effect on be-
havior and feeding times. Morphological
changes in odor receptors after oil exposure
were not detected by light and electron mi-
croscopy. The results indicate that small
amounts of oil mixed in seawater constitute a
bad odor in the lobster's environment, de-
pressing its appetite and chemical excitability
(Blumer etal. 1973).

Exposure of Marine Birds to Oil

Following the 1969 spill of 650,000 to
700,000 liters of No. 2 fuel oil into Buzzards
Bay and the adjacent Wild Harbor Marsh
near West Falmouth, Massachusetts, essen-
tially all the marsh organisms living in the
contaminated area were affected; they all ac-
cumulated oil hydrocarbons in their tissues.
Two processes apparently occur as the oil
passes through the marsh ecosystem: a pro-
gressive loss in the straight chain hydrocar-
bons in relation to the branched chain, cyclic,

and aromatic hydrocarbons, and a greater re-
tention of the higher-boiling fractions of the
contaminants by organisms higher in the food
chain (Burns and Teal 1971).

Although its feathers were not oiled, a
herring gull {Larus argentatus) that was col-
lected in Wild Harbor had substantial
amounts (584 ppm) of the whole spectrum of
fuel oil hydrocarbons in its muscle but con-
tained mostly those with straight and slightly
branched chains. The brain of this bird also
contained high residues (535 ppm), but with a
higher proportion of the aromatic compounds.
Another herring gull, collected outside the
spill area, had much lower oil hydrocarbon
residues in its muscle (10 ppm) and brain
(15 ppm), and the aromatic compounds were
not detected in the brain (Burns and Teal
1971).

Three birds that died in the 1971 San Fran-
cisco Bay oil spill contained very high resi-
dues of oil hydrocarbons in their tissues. A
composite sample of liver, kidney, brain, fat,
and heart tissue of a common murre (Uria
aalge) contained 8,820 ppm, composite liver
and kidney tissue of a surf scoter (Melanitta
perspicillata) contained 1,250 ppm, and the
liver of a western grebe (Aechmophorus occi-
dentalis) contained 9,100 ppm oil hydrocar-
bons. The composite tissues (liver, kidney,
brain, fat, and muscle) of a murre that had not
been exposed to the oil spill contained no de-
tectable oil hydrocarbons (Snyder et al. 1973).

Body fat of herring gulls breeding on Lake
Ontario in 1973 contained a number of aro-
matic hydrocarbons including several polynu-
clear aromatics. Naphthalene, 2-methyl-naph-
thalene, acetonaphthalene, and biphenyl were
identified from their retention times (Fox et
al. 1975). The sources of these aromatic hydro-
carbons remain undetermined. Accumulation
in the food chain from water or sediments
through fish is probable, but these com-
pounds, which presumably are of petroleum
origin, may have been ingested at garbage
dumps. Thus, it appears likely that aquatic
birds living in oil-polluted environments may
accumulate residues of the relatively more
persistent compounds.

Biological Effects of Oil
on Marine Birds

Aside from the reports on mortality and re-
habilitation of oiled birds, the biological ef-

fects of oil on marine birds are little known.
Important biological effects include both
acute and chronic toxicity as well as adverse
effects on physiology, reproduction, and be-
havior. Indirect effects involving the food web
and changes in habitat and food supply are
relatively unknown.

There is a distinct possibility that oil and
other environmental contaminants such as
the organochlorine compounds may act syner-
gistically (Farrington 1977).

Circumstantial evidence suggests that oil
pollution has seriously reduced populations of
certain species of marine birds in some areas
(Tuck and Livingston 1959; Tuck 1960;
Hawkes 1961; Buck and Harrison 1967; Par-
slow 1967, 1970; Bourne 1968; Clark 1973).

An oil spill can have significant effects on
populations of marine birds such as the alcids,
which often are numerous among the birds
that die in spills. Although alcids are long-
lived and have few predators once they are at
sea, they often do not breed until 3 or more
years old, most lay only a single egg per
clutch, not all adults breed every year, and
they produce an average of only one chick per
five breeding adults. These species require
more than 50 years to double their population
under optimal conditions. More than half a
century would be required for a colony to re-
cover its numbers (excluding immigration) if
reduced by one-half as the result of a large oil
spill (Clark 1969).

The potential effects of oil spills on aquatic
birds and their feeding habitat on the Cana-
dian west coast were assessed by Vermeer and
Vermeer (1975). They concluded that the pres-
ent shipping of oil plus the increased tanker
traffic along the entire British Columbia coast
that is expected to be in progress in 1977 will
result in enough oil spillage to threaten the
coastal populations of seabirds with destruc-
tion.

Concentrations of seabirds will be most vul-
nerable to spills (Vermeer and Vermeer 1975).
Three major colonies along the coast of
British Columbia are the Langara Region, the
southeast coast of the Queen Charlotte
Islands, and the Scott Islands. Alcids and
storm petrels [Oceanodroma spp.) are the
most numerous seabirds along the British
Columbia coast. Alcids are among the birds
most vulnerable to oil pollution, whereas
storm petrels are less threatened by spills be-

cause they spend more time in the air and dive
only occasionally. Waterfowl, especially div-
ing ducks, will be vulnerable to spills during
the winter as they concentrate in large
numbers in estuaries and inlets along the
British Columbia coast. The large wintering
populations of ducks, geese, and grebes along
the Fraser Delta foreshore and Boundary
Bay will be vulnerable because of their near-
ness to tanker and shipping traffic. Approxi-
mately 1 million loons, shearwaters, phal-
aropes, ducks, gulls, and geese migrate north
in the spring along west Vancouver Island.
These migrants, because of their concentra-
tion in large numbers, may be very tempo-
rarily but critically vulnerable to oil pollution.
The birds most likely to be directly affected
by spills are breeding populations of alcids
and wintering diving ducks, whereas ducks,
geese, and shorebirds, which feed in the inter-
tidal zone, may be hardest hit indirectly
through destruction of their feeding habitat
(Vermeer and Vermeer 1975). Of the ducks
threatened by destruction of their feeding
habitat, sea ducks are most vulnerable be-
cause they rely most on the marine habitat for
feeding purposes.

Feather-oiling

Large numbers of marine birds die each
year as a result of oil spills. Estimates of mor-
tality are based primarily on beach counts of
oiled birds, but such estimates may be highly
inaccurate because a significant percentage,
perhaps 50-90%, of the dead birds never wash
ashore (Clark and Kennedy 1968; Coulson et
al. 1968; Tanis and Morzer Bruyns 1968;
Hope-Jones et al. 1970).

An estimated 30,000 marine birds, of which
about 97% were common murres and razor-
bills (Alca tarda), died as a result of the Torrey
Canyon disaster (Bourne et al. 1967). Earlier,
in the winter of 1951-52, approximately
100,000 birds were lost to oil pollution on the
coasts of the British Isles (ZoBell 1962). At
least 10,000 birds, including alcids, ducks,
gulls, and kittiwakes, were killed by oil appar-
ently derived from ballast pumped from
tankers entering Cook Inlet, Alaska, during
February and March 1970 (U.S. Department
of the Interior 1970).

The population decline of murres {Uria spp.)
along the coast of Newfoundland has been as-

sociated with oil pollution (Tuck 1960), al-
though the effects probably are not related
only to those caused by feather-oiling. Numer-
ous other instances of mortality related to oil
are documented in reviews on this subject
(Clark and Kennedy 1968; Aldrich 1970; Ver-
meer and Vermeer 1974a).

The effects of oiled plumage on marine birds
vary with the properties of the oil, degree of
contamination, quantity absorbed, environ-
mental conditions, and the original condition
of the bird. Even a small patch of oil on the
feathers may mean that without care the bird
will die (Tuck 1960; Smith 1975), but in some
instances it appears that birds are able to
clean their own plumage (Phillips 1974; Smith
1975). Oiling of a bird's plumage increases me-
tabolism and causes an increased loss of body
heat to the surrounding cold water that can
readily be fatal (Lincoln 1936; Hartung 1967;
Boyle 1969; Greenwood 1970; McEwan and
Koelink 1973).

Feather-oiling appears to be a more signifi-
cant problem in cold-water areas than in areas
where water is warmer. Warm water appar-
ently causes the spilled liquid oil to form tar-
balls that are comparatively less hazardous to
birds (Bourne and Bibby 1975).

Damage to feathers may result long after
exposure and may be reflected by abnormal
wear of the plumage (Bourne 1974).

After the Torrey Canyon grounding in
March 1967, 7,849 oiled birds were captured
for cleaning and rehabilitation. One month
later, however, fewer than 6% were still alive
(Clark and Kennedy 1971).

An estimated 3,180,000 hters of bunker C
fuel oil were spilled in the massive 1971 oil
spill that occurred near the entrance to San
Francisco Bay. The California Department of
Fish and Game estimated that 7,000 aquatic
birds were exposed to the fuel oil, and more
than 4,000 of these were taken into captivity
for treatment and rehabilitation. Two weeks
after the spill, 90% of the birds had died in
spite of efforts to save them, and within 3
months mortality exceeded 96% (Orr 1971;
Snyder et al. 1973). Grebes, murres, and loons
apparently died more rapidly than the other
species affected, and ducks appeared most
hardy (Snyder et al. 1973).

Progress has since been made in the reha-
bilitation of oiled birds, and modified methods
are being used (Hay 1975). In 1973, the Inter-

national Bird Rescue Research Center treated
523 oiled birds of which 49% survived (Smith
1975).

Toxicology, Physiology, and Pathology

The great diversity of chemical compounds
in oil increases the difficulty of determining
its toxicological and physiological effects. In
addition, oil dispersants used to clean up a
spill area are also toxic and the toxicity of oil
plus dispersant usually is greater than the
toxicity of either alone (Clark and Kennedy
1968; Tarzwell 1970; Linden 1975). There also
are important species differences in suscep-
tibility (Swedmark et al. 1973).

The toxicity of some oils to ducks has been
measured under different environmental con-
ditions. Single doses of several industrial oils
produced lipid pneumonia, gastrointestinal ir-
ritiation, fatty livers, and adrenal cortical
hyperplasia. Birds that received a cutting oil
in combination with diesel oil exhibited acinar
atrophy of the pancreas. Those that received
diesel oil and a fuel oil developed toxic nephro-
sis. Cholinesterase activity was significantly
inhibited by administration of the cutting oil
and somewhat depressed by the diesel oil
(Hartung and Hunt 1966).

Ducks that had been killed by oil pollution
exhibited changes that were similar to those
in the experimentally fed birds, suggesting
that toxicity of oils is a major factor in mor-
tality of exposed birds (Hartung and Hunt
1966). Toxicity apparently is reduced through
aging of the oil because the more volatile com-
pounds are also the more toxic (Clark and
Kennedy 1968).

Birds that died after the San Francisco Bay
oil spill in 1971 were examined for pathologi-
cal changes that might have resulted from ex-
posure to oil. Intoxication from oil ingestion
appeared to be an important factor contribut-
ing to the high mortality, although the evi-
dence was circumstantial. Birds that died in
the period of high mortality had ingested oil
and exhibited dehydration, ulceration of the
intestinal mucosa, enteritis, hepatic fatty
changes, and renal tubular nephrosis (Snyder
et al. 1973). Similar pathological changes as
well as adrenal lesions and pulmonary dis-
eases have been observed in other oiled sea-
birds (Guillon 1967; Beer 1968).

Following the large 1974 oil spill in the

Straits of Magellan, a high percentage of
South American tern (Sterna hirundinacea)
chicks on an island in the spill area died
(Smithsonian Institution 1974). Although the
cause of mortality is unknown, it is possible
that the small fish that the terns ate and fed
to their young were contaminated with some
fraction of the spilled crude oil in concentra-
tions that did not harm the adults but were
toxic to the young. It is possible, however,
that the chicks died of starvation after the
adults were killed or were unable to catch
enough food for the young.

The rates at which water and sodium are
transported across the intestinal mucosa in-
crease when Pekin ducklings (Anas platyrhyn-
chos) are transferred from fresh water to a
diet containing hypertonic saline drinking
water (Crocker et al. 1974). These rate in-
creases seem to be essential for the successful
adaptation of ducklings to saline drinking
water. Ducklings given a single oral dose of a
crude oil (0.2 ml) at the start of maintenance
on saline drinking water did not develop the
characteristic rate increases. In addition, high
mucosal transfer rates that had been de-
veloped in ducklings fed saline water for
3 days ceased 24 h after they received crude
oil. Although commercial dispersant (5 ppm
or 20 ppm) in fresh or saline drinking water
had no effect on ducklings, the presence of dis-
persed crude oil (12.5-50.0 ppm) in the water
prevented the development of high mucosal
transfer rates in the ducklings given saline
water.

A reduction of the mucosal transfer rates in
seawater-adapted ducklings, through the
action of ingested crude oil, may limit the
amount of free water available to the body
(Crocker et al. 1974). Although the high mor-
tality among oil-contaminated seabirds may
be due to a variety of pathological conditions,
dehydration resulting from impairment of
mucosal transfer mechanisms may be an im-
portant factor contributing to their death.

Crude oils from eight different geographical
locations reduced the rates of sodium and
water transfer across the intestinal mucosa of
Pekin ducklings to different degrees (Crocker
et al. 1975). Administration of Kuwait crude
oil caused the greatest degree of inhibition,
and North Slope, Alaska, crude oil caused the
smallest.

Distillation fractions derived from two
chemically different crude oils were adminis-

tered to ducklings in volumes that corre-
sponded to their relative abundance in the
crude oil from which they were derived
(Crocker et al. 1975). The greatest inhibitory
effect on mucosal transfer was not associated
with the same distillation fractions from each
oil. A highly naphthenic crude oil from the
San Joaquin Valley, California, showed the
greatest inhibitory activity in the least abun-
dant (2%), low boiling point ( < 245 C) fraction.
The most abundant (47%), highest boiling
point (>482 C) fraction showed the least in-
hibitory activity. In contrast, a highly paraf-
finic crude oil from Paradox Basin, Utah,
showed the greatest inhibitory effect with the
highest boiling point fraction and a minimal
effect with the lowest boiling point fraction.
The relative abundances of these two frac-
tions in the Paradox Basin crude oil repre-
sented 27 and 28%.

Mucosal transfer inhibition by water-sol-
uble extracts of San Joaquin Valley and Para-
dox Basin crude oils was roughly proportional
to the inhibitory potency of the low boiling
point fraction of the oil (Crocker et al. 1975).
Weathered samples of these oils showed
greater effects than corresponding samples of
unweathered oils even though most of the low
molecular weight material from both oils was
either evaporated or soiubilized in the under-
lying water during the 36-h weathering
period.

Reproduction

During the nesting season, small amounts
of oil on the plumage of birds can have very
serious effects on reproduction. The oil com-
pounds that are involved, however, are essen-
tially unknown and no extensive tests have
been reported.^

Oil washed ashore on a small island in West
Germany where terns (chiefly Thalasseus
sandvicensis and Sterna hirundo) and Euro-
pean oystercatchers (Haematopus ostralegus)
were nesting. During copulation, many of the
adult terns became dorsally smeared with oil
from their mate's oiled feet, but no direct
losses among adult terns were attributed to
the oil. More than 70% of the young terns

^Reproductive effects have, however, been
studied since this manuscript was written (see
Szaro 1977).

were contaminated with oil and many of them
were unable to fly. Some of the eggs laid along
the high-tide mark failed to hatch after they
became contaminated with oil (Rittinghaus
1956).

After ingestion of a relatively nontoxic lu-
bricating oil (2 g/kg), one mallard {Anas platy-
rhynchos) and two Pekin ducks stopped lay-
ing for about 2 weeks. Very small quantities
of oil coated on mallard eggs reduced their
hatchabihty to 21%, compared with 80% for
unoiled eggs. Experimentally oiled mallards
continued to incubate their clutches, but their
eggs failed to hatch (Hartung 1965).

In an experimental application to test the
effects of 2,4-D and diesel fuel on eggs of ring-
necked pheasants {Phasianus colchicus), there
was no adverse effect by the 2,4-D on hatch-
ability, but apphcation of the diesel fuel re-
duced hatchability to zero (Kopischke 1972).

Behavior

Exposure to oil causes some obvious
changes in behavior patterns of birds because
they abandon all activities to attempt to clean
the oil from their feathers by preening (Smith
1975). There may be other serious but less
readily observed direct effects that influence
the birds' ability to locate food, to migrate, or
to perform other essential activities.

As discussed earlier, small amounts of oil in
the water cause significant changes in be-
havior of certain marine organisms. Modified
behavior among any of the numerous species
of animals in the food webs may have serious
indirect implications for the welfare of marine
birds that depend upon them,

Organochlorines

By 1971, and perhaps earlier, it became un-
likely that any bird dependent upon marine
food webs anywhere in the world was free of
contamination by the synthetic organo-
chlorine compounds that have become ubi-
quitous pollutants in the global ecosystems
(Sladen et al. 1966; Risebrough and Berger
1971; Bogan and Bourne 1972; Bourne and
Bogan 1972; Bennington et al. 1975; Rise-
brough 1977; Walker 1977; White and Rise-
brough 1977). More information is available
on the global distribution patterns of organo-

chlorines than for other chemicals in marine
birds. Several direct biological effects of or-
ganochlorines on marine birds are known.
Other relevant information is available on the
distribution of these pollutants in estuarine,
freshwater, and terrestrial ecosystems, as
well as their biological effects on other birds.

The most abundant synthetic organo-
chlorine compound in tissues and eggs of ma-
rine birds is frequently p,p '-DDE, a derivative
of p,p'-DDT, which is the principal component
of the commercial insecticidal mixture (Rise-
brough et al. 1968; Jensen et al. 1969; Koeman
et al. 1969). Other DDT compounds fre-
quently present in marine birds arep,p'-DDD,
p,p'-DDT, and o,p'-DDT (Bennington et al.
1975).

Polychlorinated biphenyls (PCB's), or
chlorobiphenyls, consist of a mixture of com-
pounds differing in chlorine content and the
position of chlorine atoms on the parent bi-
phenyl molecule. Pentachlorobiphenyls and
hexachlorobiphenyls usually constitute the
majority of the chlorobiphenyls present in
marine birds, but trichlorobiphenyls and tet-
rachlorobiphenyls are occasionally present
(Risebrough and de Lappe 1972; White and
Risebrough 1977).

A number of other synthetic organochlorine
compounds have been detected in marine
birds, but almost always at levels substan-
tially lower than those of the DDT and PCB
compounds. Hexachlorobenzene (HCB), which
has been found in tissues of great cormorants
(Phalacrocorax carbo), sandwich terns
(Thalasseus sandvicensis), and common eiders
(Somateria mollissima) from coastal areas of
the Netherlands (Koeman and van Genderen
1972; Koeman et al. 1972a), has been consid-
ered a potentially hazardous marine pollutant
(National Academy of Sciences 1975b). The
HCB is used as a fungicide but may enter the
marine environment in significant quantities
as a component of the tarry waste products
from the manufacture of chlorinated hydro-
carbons such as perchloroethylene and carbon
tetrachloride that are frequently discharged
at sea (Environmental Protection Agency
1973).

Chlorinated styrenes were identified by gas
chromatography/mass spectrometry in tis-
sues of common eiders, sandwich terns, and
great cormorants from the Netherlands (Ten
Noever de Brauw and Koeman 1972) and in

tissues of great blue herons {Ardea herodias)
from Lake St. Clair, Michigan (Reichel et al.
1977); their source apparently remains un-
known. Chlorinated naphthalenes also were
identified in tissues of the great cormorant
(Koemanetal. 1973).

Mirex was measured in eggs of herons and
white ibis {Eudocimus albus) from estuarines
of the U.S. Atlantic and Gulf coasts (Ohlen-
dorf et al. 1974) and also in the blubber of a
seal (Phoca vitulina) from the Netherlands
(Ten Noever de Brauw et al. 1973). In addition
to its use as an insecticide, mirex, under
various trade names, is used as a flame
retardant.

Dieldrin was found in eggs and tissues of
several species of marine birds inhabiting
coastal waters of Great Britain (Robinson et
al. 1967) and of New Zealand, and also in
pelagic species such as the sooty shearwater
(Puffinus griseus) breeding in sub-Antarctic
islands of New Zealand (Bennington et al.
1975). It is accumulated by ospreys (Pandion
haliaetus) and bald eagles {Haliaeetus leuco-
cephalus) feeding on coastal marine fish of the
eastern United States (Mulhern et al. 1970;
Belisle et al. 1972; Cromartie et al. 1975; Wie-
meyer et al. 1975).

Endrin was detected in brown pelicans
(Pelecanus occidentalis) from Florida
(Schreiber and Risebrough 1972), the Gulf of
California (Risebrough et al. 1968), and
Louisiana (J. D. Newsom, personal communi-
cation). White pelicans (Pelecanus erythro-
rhynchos) from Louisiana also contained en-
drin (J. D. Newsom, personal communica-
tion).

Chlordane compounds, principally oxy-
chlordane and cis-chlordane, were found in
eggs of herons from the eastern U.S. estuaries
(Ohlendorf et al. 1974) and in fish and common
terns {Sterna hirundo) from Long Island
Sound (R. W. Risebrough and P. Robinson,
unpublished data).

Heptachlor epoxide, toxaphene, and the
several isomers of hexachlorocyclohexane
(benzene hexachloride or BHC) are occa-
sionally found in estuarine environments.
Heptachlor epoxide and BHC isomers have
been reported in Antarctic birds breeding in
the South Orkneys (Tatton and Ruzicka 1967)
but their identification has not been con-
firmed (Risebrough 1977).

The occurrence, distribution, and effects of

organochlorines on wildlife, principally terres-
trial, freshwater, and estuarine species, have
been summarized in other recent reviews
(Prestt and Ratcliffe 1972; L. F. Stickel 1972,
1973; Blus et al. 1977b; W. H. Stickel 1975;
Ketchum et al. 1975; L. F. Stickel and F. E.
Hester, unpublished manuscript). Earlier
studies of the transport of PCB's to the ma-
rine environment also have been reviewed
(Nisbet and Sarofim 1972; Panel on Hazar-
dous Trace Substances 1972). More recently
the environmental effects of PCB's (Peakall
1975), their chemical properties (Hutzinger et
al. 1974), and the transfer of organochlorine
compounds to the marine environment and
their incorporation into marine food webs
have been reviewed (Risebrough et al. 1976a).

Although levels of organochlorine com-
pounds other than those of the DDT and PCB
groups may occasionally be present at levels
deleterious to birds in estuaries, levels in the
offshore marine environment are usually well
below those considered harmful to marine
birds. In the present review, principal em-
phasis will therefore be placed on the DDT
and PCB compounds.

Exposure of Marine Birds
to Organochlorines

Organochlorine residue data are available
from coastal regions, but there have been rela-
tively few studies of chlorinated hydrocarbon
contamination of marine birds in areas that
are far from known pollution sources.

All eggs of the Adelie penguin {Pygoscelis
adeliae) from widely separated localities in the
Antarctic (Risebrough and Carmignani 1972;
Risebrough 1977), eggs and tissues of birds
from the Aleutians (White and Risebrough
1977) and from sub-Antarctic areas of New
Zealand (Bennington et al. 1975), and tissues
of birds from the eastern North Atlantic
(Bourne and Bogan 1972; Bogan and Bourne
1972) contained residues of DDT compounds.
All samples also contained detectable levels of
chlorobiphenyl compounds, frequently at
levels higher than the total concentration of
the DDT group. The DDT and chlorobiphenyl
compounds also were detected in all samples
obtained from remote terrestrial and fresh-
water Arctic ecosystems (Risebrough and
Berger 1971; Walker 1977).

Although the data are few and sample sizes

Table l.Mean PCB and DDE residues (ppm lipid weight) in cormorants
(Phalacrocorax spp. )

Percent

Locality, date

Species

Tissue

lipid

PCB's

DDE

PCB/DDE

Amchitka, 1971*

Red-faced cormorant
(P. unle)

Yolk

20.0

19.0

3.8

5.0

Amchitka, 1974°

Red-faced cormorant

Pectoral muscle

—

21.0

3.5

6.0

Agattu, 1974°

Red-faced cormorant

Pectoral muscle

3.8

14.0

2.4

6.0

Amchitka, 1974°

Pelagic cormorant
(P. pelagicus)

Pectoral muscle

4.0

8.0

0.8

10.0

Auckland Islands,

Auckland Island

Egg lipid

100.0

0.3

0.9

0.3

1972b

shag

Iceland, 1973'^

Shag

(P. aristotelis)

Egg

5.0

23.0

3.8

6.0

Great cormorant

Egg

4.8

10.0

3.0

Peru, 1969*^

Guanay

(P. bougainuillei)

Egg lipid

100.0

15.0

12.2

1.2

Southern Cah-

Double-crested

Egg lipid

100.0

87.0

754.0

0.1

fornia, 1969e

cormorant

Greenland, 1972*

Great cormorant

Body fat

23.0

9.8

2.3

°White and Risebrough 1977.

^Bennington et al. 1975.

'^J. A. Sproul et al., unpublished manuscript.

'^R. W. Risebrough et al., unpubhshed manuscript.

^Gressetal. 1973.

^Braestrup et al. 1974.

are frequently small, a general picture of
global marine contamination by DDT and
PCB compounds can be presented. Some of
the available data on cormorants have been
summarized (Table 1). With the exception of
the residue values reported for the double-
crested cormorant (Phalacrocorax auritus) in
southern Cahfornia, where the birds were ex-
posed to industrial contamination from the
Los Angeles area (Gress et al. 1973), the
samples were obtained from areas reasonably
remote from point sources of contamination.
The DDE residues in the Auckland Island
shags (Phalacrocorax carunculatus) were
somewhat lower than in cormorants from
Amchitka and Agattu at the equivalent lati-
tude in the northern hemisphere. However,
PCB values in the southern hemisphere birds
were lower by 1-2 orders of magnitude.

Other data from biocenotic equivalents in
the two areas support the conclusion that
DDE levels are slightly lower in the southern
than in the northern hemisphere but that PCB
values are lower by 1-2 orders of magnitude

(Bennington et al. 1975; White and Rise-
brough 1977). The DDE levels in an egg of a
New Zealand falcon (Falco novae seelandiae)
were equivalent to those in eggs of peregrines
from Amchitka, but PCB levels were an order
of magnitude lower. Comparable differences
were found between auklets (Aethia pusilla
and Cyclorrhynchus psittacula) and the tufted
puffin (Lunda cirrhata) of the Aleutians and
the diving petrel (Pelecanoides urinatrix) from
the Snares Islands of southern New Zealand.

The DDE levels in eggs of the guanay were
somewhat higher than those from New Zea-
land or the Aleutians, suggesting local
sources of DDT contamination in Peru (R. W.
Risebrough et al., unpublished manuscript).

A comparison of the cormorant samples
from the Aleutians and the eggs of two cor-
morants (Phalacrocorax carbo and P. aris-
totelis) breeding in Iceland suggests that
levels of DDE and PCB contamination in the
two oceanic areas are similar. In five species
of fish obtained from Amchitka in 1974, DDE
residues ranged from 1 to 5 ppb; PCB resi-

dues ranged from 8 to 20 ppb (White and
Risebrough 1977). Residues of DDE in seven
species of fish obtained from the coastal
waters of Iceland in 1973 ranged from 1 to
9 ppb; PCB levels ranged from 8 to 20 ppb
(J. A. Sproul et al., unpublished manuscript).
On a lipid basis, PCB residues expressed as
tri-, tetra-, or penta-chlorobiphenyls ranged
from 0.3 to 2 ppm in the Amchitka fish and
from 0.2 to 3 ppm in the Icelandic fish. Body
fat of great cormorants from Greenland
(Braestrup et al. 1974) contained comparable
PCB levels and somewhat higher DDE levels
than great cormorants from Iceland.

A comparison of DDE and PCB residue
levels in black-legged kittiwakes {Rissa tridac-
tyla), fulmars (Fulmarus glacialis), and thick-
billed murres {Uria lomvia) from Amchitka
and Iceland suggests somewhat higher levels
in the Icelandic birds, although residues are of
the same order of magnitude and with com-
parable ratios (J. A. Sproul et al., unpubhshed
manuscript; White and Risebrough 1977). The
differences may reflect a higher level of con-
tamination in those areas of the ocean where
the Atlantic birds spend the winter months.

Earlier data (Risebrough et al. 1968) sug-
gested that DDT compounds were more abun-
dant than PCB's in Pacific waters. However,
many of the samples were from coastal CaU-
fornia waters where DDT contamination was
particularly severe.

Residue levels and PCB:DDE ratios in the
breast muscles of Icelandic birds obtained in
1973 were comparable to those in birds ob-
tained earlier from areas north of Britain, in-
dicating that no decline in residue concentra-
tions in birds had occurred over that short
interval (Bourne and Bogan 1972; J. A.
Sproul et al., unpublished manuscript).

From the Pacific, the visceral fat of 7 black-
footed albatrosses {Diomedea nigripes) and
22 Laysan albatrosses (Z). immutablis) from
Midway Island contained mean DDE levels of
22 and 8 ppm, and mean PCB levels of 14 and
2 ppm (Fisher 1973). These species are re-
stricted to the North Pacific, usually about
20° N and feed primarily on squid. In Hawaii,
DDE concentrations in four eggs of the dark-
rumped petrel (Pterodroma phaeopygia)
ranged from 0.07 to 1.14 ppm (0.6-11.5 ppm,
lipid weight) (King and Lincer 1973). PCB
values were not reported.

From the tropical Atlantic, DDE and PCB

levels in the breast muscle of 28 adult sooty
terns (Sterna fuscata), breeding on the Dry
Tortugas, were 2.5 and 7.8 ppm (lipid weight);
mean percentage of hpid was 2.6% (P. G.
Connors et al., unpublished manuscript).

In these areas of the Atlantic and Pacific,
comparatively remote from sources of con-
tamination, PCB residue concentrations gen-
erally exceeded those of the DDT compounds.
In the New Zealand (including the sub-Ant-
arctic islands) samples, however, DDT resi-
dues were frequently present at higher con-
centrations than the sum of PCB's (Benning-
ton et al. 1975).

In the Antarctic, few eggs of the Adelie pen-
guin obtained in 1970 or earlier from widely
separated locaUties contained PCB's at de-
tectable levels (Risebrough et al. 1976a).
Maximum amounts of PCB's in eggs of the
Adelie penguin obtained from Cape Crozier in
October 1967 were less than one-eighteenth of
the concentration of DDT compounds (Rise-
brough et al. 1968). Subsequent analysis of
some of these eggs revealed the presence of
PCB. PCB's were detected also in eggs of the
Adelie penguin, chinstrap penguin (Pygo-
scelis antarctica), and the gentoo penguin (P.
papua) obtained in the Antarctic Peninsula in
1975, although at concentrations less than
those of the DDT compounds (Risebrough et
al. 1976b). The preponderance of DDT com-
pounds in the Antarctic, the most remote area
receiving chlorinated hydrocarbons from at-
mospheric or oceanic transport, apparently
reflects the relative use of these two groups of
compounds in the southern hemisphere.

In coastal areas local conditions usually de-
termine the contamination patterns. For
example, liquid chemical wastes discharged
by insecticide manufacturing plants in CaU-
fornia and the Netherlands subsequently en-
tered the sea and caused significant organo-
chlorine contamination of coastal birds. High
DDT concentrations were found in northern
anchovies (Engraulis mordax) from Los An-
geles Harbor in 1965 (Risebrough et al. 1967).
Subsequent investigations documented ex-
ceptionally high levels of DDT compounds in
the coastal birds, including the brown peli-
cans (Risebrough 1972; Anderson et al. 1975)
and double-crested cormorants (Gress et al.
1973). When the company began to dispose of
its hquid wastes in a sanitary landfill in 1970,
input of DDT compounds into the sea began

to decline (Carry and Redner 1970; Redner
and Payne 1971; D. R. Young et al., unpub-
lished manuscript); residues in fish and in the
brown pelicans also began to decline (Ander-
son etal. 1975).

In 1964, sandwich terns and spoonbills
(Platalea leucorodia) were found dying on the
island of Texel in the Dutch Wadden Sea. The
birds were in tremors and convulsions, signs
comparable to those found in other birds
poisoned by organochlorine insecticides (Koe-
man and van Genderen 1965, 1966). Studies of
the distribution of chlorinated hydrocarbons
in birds, fish, and mussels (Mytilus edulis)
from localities along the Dutch and West
German coasts and in the eggs of seabirds of
Great Britain indicated a point source of con-
tamination by dieldrin, endrin, and telodrin.
Telodrin, an insecticide not used in Europe at
that time, was being manufactured with diel-
drin and endrin in a factory near the mouth of
the Rhine River. When it was discovered that
these residues were coming from the insecti-
cide plant, measures were taken to eliminate
discharge; residue levels in the local seabirds
began to decline and the sharp decrease in
population numbers was halted (Koeman et
al. 1968).

In addition to these two incidences, coastal
contamination from local but diffuse sources
has resulted in high levels of organochlorines
in birds in Japan, North America, and
Europe. Levels of PCB in Japanese birds, in-
cluding several species of gulls, were compa-
rable to those in industrial areas of North
America and Europe (Fujiwara 1974). The
PCB residues in breast muscle of eight little
egrets (Egretta garzetta) that were found
dead or dying in Tokyo Bay ranged from
0.3 to 180 ppm (22-1,600 ppm, lipid basis)
with a geometric mean of 9 ppm. Residues of
PCB in the breast muscle of eight black-tailed
gulls (Larus crassirostris) ranged from
3-39 ppm, with a geometric mean of 13 ppm
(Doguchi 1973).

In western North America comparatively
high levels of DDT and PCB contamination
were found in common murres (Gress et al.
1971) and the ashy storm petrels (Oceano-
droma homochroa) (Coulter and Risebrough
1973) breeding on the Farallon Islands and in
great egrets {Casmerodius albus) and great
blue herons breeding at a coastal site (Faber
et al. 1972) near local sources of pollution.

Most eggs of marine birds from the Strait of
Georgia contained more PCB's and DDE and
had a higher PCB:DDE ratio than did eggs
from the west coast of Vancouver Island and
from the Queen Charlotte Islands (Table 2).
This comparison within a relatively small re-
gion (i.e., the Pacific Coast of British Colum-
bia) further illustrates the principle that eggs
from birds nesting farther at sea are likely to
contain lower levels of organochlorines than
those nesting nearer the mainland. Average
PCB levels in these samples almost always ex-
ceeded those of DDE.

Fish, crabs, and shellfish were collected
from the lower Eraser River, its estuary, and
selected areas of Georgia Strait (Albright et
al. 1975). Generally, PCB's were present at
higher levels than DDE, and greatest concen-
trations of these compounds occurred in biota
from waters adjacent to the city of Van-
couver. With one exception, animals from
Georgia Strait and those away from the im-
mediate influence of Eraser River water con-
tained no detectable levels of chlorinated
hydrocarbons.

High levels of DDE and PCB in double-
crested cormorants from the Bay of Fundy
(Zitko and Choi 1972; Zitko et al. 1972) most
likely have resulted from past DDT use in
New Brunswick and from diffuse sources of
PCB's along the eastern North American
coastline. Similarly, contamination levels in
ospreys (Wiemeyer et al. 1975; Spitzer et al.
1977) in coastal Connecticut, Massachusetts,
New York, and New Jersey most likely were
derived from local sources of contamination.

Bald eagles found sick or dead in the United
States during 1966-72 were analyzed for or-
ganochlorines (Mulhern et al. 1970; BeHsle et
al. 1972; Cromartie et al. 1975); DDE, DDD,
dieldrin, and PCB's were detected in most of
the 145 eagle carcasses. Eighteen of the
eagles contained possibly lethal levels
(greater than 4 ppm) of dieldrin. Since 1964
when data were first collected, 8 of the 17
eagles obtained from Maryland, Virginia,
South Carohna, and Florida possibly died
from dieldrin poisoning. All four specimens
from Maryland and Virginia were from the
Chesapeake Bay Tidewater area.

In December 1973, eight ruddy ducks
(Oxyura jamaicensis) killed in an oil spill on
the Delaware River (White and Kaiser 1976),
contained DDE (1.1-4.5 ppm) and PCB's (2.8-
10 ppm). Levels of DDT and DDD were below

Table 2. Mean PCB and DDE residues (ppm lipid weight) in seabird eggs from the Strait of
Georgia, the west coast of Vancouver Island, and the Queen Charlotte Islands in British Col-
umbia, 1970 (K. Vermeer, unpublished data).

Locality

Species

Percent
N lipid PCB's DDE PCB/DDE

Strait of Georgia

Mandarte Island

Mittlenatch Island

Vancouver Island

Cleland Island

Queen Charlotte Islands
Skedans Island

Lucy Island

Northwest Rocks

Double-crested cormorant

Pelagic cormorant

Glaucous-winged gull
{Larus glaucescens)

Pelagic cormorant

Pigeon guillemot
{Cepphus columba)

Glaucous-winged gull

Leach's petrel
(Oceanodroma leucorhoa)

Pigeon guillemot

Tufted puffin
Glaucous-winged gull

Fork-tailed petrel
(O. furcata)

Pigeon guillemot
Glaucous-winged gull

Rhinoceros auklet

(Cerorhinca

monocerata)

Glaucous-winged gull

6.9

207.0

59.0

3.5

5.3

50.0

15.0

3.3

6.0

41.5

12.5

3.3

4.4

122.0

12.0

10.2

10.5

34.0

6.0

5.7

8.0

19.0

6.0

3.2

12.7

8.5

17.0

0.5

10.8

24.0

12.0

2.0

10.0

6.5

4.0

1.6

8.6

30.0

18.5

1.6

2
10

10
3

29.6

11.7
8.3

15.0

9.7

8.7

51.0

3.6
6.0

13.0

6.0
4.2

14.0

1.4
4.0

18.0

3.0
2.6

3.6

2.6
1.5

0.7

2.0
1.6

0.34 ppm in all but one sample. Dieldrin and
HCB were present in seven samples, but
neither exceeded 0.36 ppm.

In a survey of organochlorine residues in 21
aquatic bird species at 31 locations in Alberta,
Saskatchewan, and Manitoba, DDE and diel-
drin levels were higher in eggs of larids and
fish-eating birds than in those of geese and
ducks, presumably reflecting different trophic
levels between those two groups of birds (Ver-
meer and Reynolds 1970).

On the Niagara Peninsula, an area of On-
tario that is intensively developed for agricul-
ture and heavy industry and has a large urban
population, eggs were collected in 1972 from
20 species of birds having a variety of feeding
habits (Frank et al. 1975). Representative
species were obtained from both the terres-

trial and aquatic food chains. Highest total
DDT residues were in the eggs of aquatic car-
nivores, including common tern (22.4 ppm),
herring gull (10.4 ppm), black-crowned night
heron (Nycticorax nycticorax; 7.8 ppm), and
black tern (Chlidonias niger; 7.6 ppm). Herbi-
vores and insectivores contained lower total
DDT residues regardless of the environment
in which they fed. The highest mean residues
of PCB's also were in carnivores in the aquatic
food chain, including herring gulls (74 ppm),
common terns (42 ppm), and black-crowned
night herons (27 ppm).

Eggs of anhingas {Anhinga anhinga),
herons, and ibises were collected during the
1972 nesting season at coastal and inland lo-
caUties from Florida to New Jersey (Ohlen-
dorf et al. 1974). Measurable residues of DDE

occurred in all 209 eggs. The highest mean
value (4.0 ppm) was found in great egrets
from New Jersey. Among the coastal locali-
ties, levels of DDE as well as total DDT pro-
gressively declined toward the south. The
PCB's occurred second most frequently and
also reached their highest mean level
(4.2 ppm) in the great egret eggs from New
Jersey. Other pollutants occurred less fre-
quently and at lower levels.

In Great Britain, where the presence of or-
ganochlorine pollutants in seabird eggs was
first demonstrated (Moore and Tatton 1965),
organochlorine residues in seabird eggs from
a number of colonies have been monitored.
Populations of common puffin (Fratercula arc-
tica) in Great Britain declined (Flegg 1971,
1972), but those birds analyzed have not
shown excessively high contamination levels.
Birds from Saint Kilda contained 7.6 ppm of
PCB (61 ppm in fat), but seven other puffins
contained lower concentrations. Five eggs ob-
tained from Saint Kilda in 1969 contained
lower residues of PCB's, DDE, and dieldrin
than did eggs of either the common murre or
the razorbill (Parslow et al. 1972). figgs of the
murre from Lundy, Skomer, and Berry Head
contained lower PCB levels than did eggs of
the kittiwakes from the same location, but
DDE levels were lower in the kittiwakes
(Parslow 1973). In some localities on the Brit-
ish coast, eggs of murres contained levels of
PCB's that were as high as those reported
from California and the Baltic, but DDE
levels were lowest in Britain.

Biological Effects of
Organochlorines on Marine Birds

Although there is a considerable amount of
information on residue concentration and re-
productive effects of organochlorines in ma-
rine birds, there is relatively little information
on toxicology, physiology, and pathology in
these species. Therefore, it is particularly rele-
vant to consider also such effects in the more
frequently studied terrestrial species.

Toxicology, Physiology, and Pathology

Evidence is substantial that PCB's may
have contributed to the mortality of contami-
nated birds. Great cormorants found dead in

the Netherlands may have died of PCB
poisoning (Koeman et al. 1973). Residues in
the brain and liver were equivalent to those in
birds poisoned through feeding of the PCB
preparation Clophen A60. Chlorinated diben-
zofurans, however, were present in the com-
mercial PCB mixture (Vos et al. 1970) and
may have contributed to the mortality of the
experimental birds. Therefore, the residue
levels in tissues may not be equivalent in the
toxicological sense.

The occasional "wrecks" of seabirds, par-
ticularly of common murres, are usually asso-
ciated with storms. In 1970 more than
100,000 murres died in Bristol Bay, Alaska,
following stormy weather (Bailey and Daven-
port 1972). There had been no oil spills in the
area. The birds were emaciated and appar-
ently had starved as a result of an inability to
find food during the prolonged storm, but
they were not analyzed for organochlorines.

In Great Britain, PCB concentrations in the
livers of gannets {Moms bassanus) that died
during large-scale mortality incidents in 1972
ranged from 3,300-9,600 ppm, lipid weight;
DDE concentrations ranged from 260-
520 ppm, lipid weight (Parslow et al. 1973).
Organochlorine concentrations of this magni-
tude might contribute to the death of marine
birds either through direct poisoning follow-
ing mobilization of fat or through more subtle
sublethal effects on the birds at a time of en-
vironmental stress.

Because concentrations of chemicals in the
body are greatly affected by weight gains and
losses, it is sometimes more useful to compare
total body loads. Estimated body contents of
PCB's and DDE in five murres found dead
during a 1969 wreck in the Irish Sea were
2,700 /ig (range 800-8,900) and 673 /ig (314-
1,535) (Holdgate 1971). Five birds that were
shot in the same general area had 3,500 (ig
(800-7,200) of PCB's and 1,484 ng (468-3,211)
of DDE. However, eight other murres that
died in the wreck had an average estimated
body burden of 4,660 ng of PCB, twice as
much as in nine other apparently healthy
birds that were collected (Parslow and Jef-
feries 1973). Depletion of body fat during
times of hunger could be expected to mobilize
chlorinated hydrocarbons, providing addi-
tional stress when the birds are poorly
equipped to cope with it. The overall contribu-
tion of chlorinated hydrocarbons, particularly

PCB's, to such mortality remains to be
determined.

A glaucous gull {Larus hyperboreus) found
in convulsions on Bear Island in the Arctic
contained 311 ppm of PCB's and 67 ppm
DDE in the liver (Bogan and Bourne 1972); its
weakened condition and abnormal coordina-
tion were attributed to these high levels. The
glaucous gulls in this colony were feeding on
the eggs of other seabirds.

Necropsy findings and the high level of
DDD (200 ppm) in the brain of a common loon
{Gavia immer) found in a soybean field in
Madison County, Mississippi, indicate that
the bird died of DDD poisoning (Prouty et al.
1975).

Experiments have been conducted with cap-
tive birds to determine which tissue might
contain chemical residues that are diagnostic
of organochlorine poisoning (L. F. Stickel et
al. 1966; W. H. Stickel et al. 1969, 1970, 1973;
Stickel and Stickel 1970). There is little doubt
that many closely related compounds have a
lethal additive effect in the nervous system
(Ludke 1976; J. L. Ludke and W. H. Stickel,
personal communication). Chemical residues
in the brain, in association with pathological
conditions of the body, may reveal that the
compounds caused death. Lethal ranges have
been established for DDT, DDD, DDE, diel-
drin, and mirex. Suggestions also have been
made for weighing and summing brain resi-
dues of DDT, DDD, and DDE for interpreta-
tion of field specimens.

Recent studies of the induction of hepatic
enzymes by PCB's have been reviewed
(Peakall 1975). Levels of PCB in many sea-
birds may be assumed to be sufficient to in-
crease the activity of various mixed function
oxidase enzymes. Elevated activity levels of
these enzymes also enhance steroid metab-
oUsm and degrade nonpolar compounds of
foreign origin. The biological consequences of
increased steroid metabolism are unknown
but birds may compensate for the higher level
of steroid metabolism by increasing levels of
synthesis.

Teratogenic effects observed in experi,-
mental feeding studies with PCB's have in-
cluded malformations of the eye, legs, and
beaks (Carlson and Duby 1973; Tumasonis et
al. 1973; Cecil et al. 1974). Such abnormalities
may have been caused by contaminant di-
benzofurans in the PCB mixtures (Vos et al.

1970; Bowes et al. 1975). Similar abnormali-
ties have been found in common terns breed-
ing in Long Island Sound (Hays and Rise-
brough 1972) but the cause remains unknown;
a Unk with PCB or chlorinated dibenzofurans
has not yet been proven.

Diets containing 10 and 30 ppm (dry
weight) DDE were fed to black ducks (Anas
rubripes), and diets containing 1, 5, and
10 ppm (dry weight) DDE were fed to mal-
lards (Longcore et al. 1971a). Among the
results were the following changes in black
duck eggshell composition: (1) significant in-
crease in the percentage of magnesium, (2) sig-
nificant decreases in barium and strontium,
(3) increases (which approached significance)
in average percentage of eggshell sodium and
copper, (4) a decrease in shell calcium that ap-
proached significance, (5) patterns of mineral
correlations that in some instances were dis-
tinct to dosage groups, and (6) inverse correla-
tions in the control group between eggshell
thickness, magnesium, and sodium.

Changes in mallard eggshells were: (1) signi-
ficant increase in percentage of magnesium at
5 and 10 ppm DDE, (2) significant decrease in
aluminum at 5 and 10 ppm DDE, (3) a signifi-
cant decrease of calcium in the 10 ppm DDE
group, and (4) an increase in average per-
centage of sodium in eggshells from DDE-
dosed ducks that approached significance.

Blood samples were taken for 2 successive
years from canvasback ducks [Ay thy a valis-
ineria) trapped in the Chesapeake Bay (Dieter
et al. 1976). The first winter (1972-73), five
plasma enzymes known to respond to organo-
chlorine poisoning were examined. Altera-
tions in enzyme activity indicated tissue dam-
age (specifically in membrane permeability) at
the cellular level. Abnormal enzyme eleva-
tions suggested that 20% of the population
sampled (23 of 115 ducks) might contain ele-
vated levels of organochlorine contaminants,
but no residue analyses were performed. The
second winter (1973-74), two of the same en-
zymes, aspartate aminotransferase and lac-
tate dehydrogenase, were assayed in 95 blood
samples. The PCB concentrations in represen-
tative blood samples were significantly
(P < 0.05) correlated with plasma aspartate
aminotransferase activity.

Male coturnix quail {Coturnix cotumix)
were fed diets containing graded levels of
DDE, PCB (Aroclor 1254), malathion, and

mercuric chloride (Dieter 1974). At 12 weeks,
increases in each of the activities of five
plasma enzymes (creatine kinase, aspartate
aminotransferase, cholinesterase, fructose-di-
phosphate aldolase, and lactate dehydroge-
nase) of birds were proportional to the log
dose of the respective agents. In addition, the
pattern of enzyme responses in the experi-
mental groups had changed, and was illustra-
tive of the specific type of substance that had
been fed. The data suggest that qualitative
and quantitative identification of environ-
mental contaminants in birds, and perhaps a
variety of wild animals, may be possible by
utilization of multiple plasma enzyme assays.
Residue analyses after 12 weeks of feeding
showed that DDE accumulated in carcasses
and Uvers at concentrations up to fourfold
higher than those in the diets. In contrast,
residues of Aroclor 1254 attained in carcasses
were identical to, and in livers one-half of, the
concentration in the feed.

Wild-trapped starlings {Sturnus vulgaris)
were fed concentrations of DDE or Aroclor
1254 (5, 25, and 100 ppm, dry weight) that
were found to be sublethal when fed to pen-
reared coturnix quail for 12 weeks (Dieter
1975). Although the experimental design had
been to compare plasma enzyme responses at
3, 7, and 12 weeks, reliable measurements
could only be made through 7 weeks of the ex-
periment because of unexpected mortality.
Variations in enzyme response were greater in
wild than in pen-reared birds, but not enough
to mask the toxicant-induced changes in en-
zyme activity. Cholinesterase, lactate dehy-
drogenase, creatine kinase, and aspartate
aminotransferase activities increased in those
fed the organochlorine compounds. Evalua-
tion of enzymatic profiles appears to be a po-
tentially valuable technique to monitor the
presence of toxicants in wild populations, es-
pecially if used to complement standard
chemical residue analyses. After feeding for
7 weeks, liver residues of either organo-
chlorine compound were about threefold
higher than the concentrations fed daily.
However, 4 times as much DDE as Aroclor
1254 had accumulated in the carcasses.

Dietary DDE at levels from 10 to
1,000 ppm (dry weight) inhibited nasal gland
secretion in mallards maintained in fresh
water (Friend et al. 1973). However, in subse-
quent studies on the effects of dietary DDE

(10-250 ppm, dry weight) on osmoregulation
and nasal gland function in mallards, Pekin
ducks, black guillemots (Cepphus grylle), and
common puffins, DDE had minimal effects on
plasma electrolyte levels and total nasal gland
Na,K-ATPase activities in each of these
species (Miller et al. 1976). Liver DDE levels
in experimental ducks and guillemots were
comparable with those reported for seabirds
found dead after kills; levels in starved puffins
were much higher. Therefore, DDE at envi-
ronmental levels may not affect osmoregula-
tion of nasal gland Na,K-ATPase in ducks or
in these two species of marine birds.

Coturnix quail were fed 1 ppm (dry weight)
dieldrin, 2 ppm DDE, or the two chemicals to-
gether (Ludke 1974). When fed alone, both
dieldrin and DDE reached their highest con-
centrations in the birds' livers after 28 days
on treatment, followed by a slight decrease
after 56 days. In whole-body samples (carcass
minus liver), dieldrin residues increased stead-
ily throughout the treatment period. Dieldrin
residues in the birds exposed to dieldrin alone
were always similar to residues in birds that
were exposed to dieldrin in combination with
DDE. In birds fed DDE, either alone or in
combination with dieldrin, DDE residues in
the carcass increased similarly for 28 days.
After 56 days, DDE residues were signifi-
cantly greater in the birds fed the dieldrin and
DDE mixture. The continued increase of DDE
residues when both DDE and dieldrin were
fed suggests an interaction in which dieldrin
promotes an increased uptake or retention of
DDE.

No weight loss or mortality occurred among
bobwhite (Colinus virginianus) fed a control
diet or those fed chlordane (10 ppm, dry
weight) alone. However, birds that were
fed endrin (10 ppm, dry weight) or a combina-
tion of chlordane and endrin lost weight and
died within a few days (Ludke 1976). Mori-
bund individuals had lost considerable body
weight and contained much less body fat than
did individuals that were not exhibiting signs
of intoxication when sacrificed. Birds that
died from intoxication averaged weight losses
of 32.2% (endrin-treated) and 31.4% (chlor-
dane + endrin-treated) when compared with
the control group. Individuals that survived
exposure had significantly lower brain resi-
dues than those that died. Residues of endrin
were significantly lower (by 38%) in brains of

birds that died from endrin plus chlordane
than in those dying from endrin alone. These
data indicate that closely related toxicants
may have an accumulative effect at the site of
action.

Two of 14 male American kestrels {Falco
sparuerius) died after 14 and 16 months on a
diet containing 2.8 ppm DDE (Porter and
Wiemeyer 1972). The brains of the two birds
contained DDE residues of 213 and 301 ppm
compared with an average of 14.9 ppm (range,
4.5-26.6 ppm) for 11 of the adult males that
were sacrificed after 12 to 16 months on
dosage. Each of the two birds that died had
lost about one-third of its weight since treat-
ment began and necropsy revealed typical
characteristics (reduced pectoral muscle and
badly depleted fat reserves) of organochlorine
poisoning.

Endrin was consistently the most toxic of
89 pesticidal chemicals that were tested for
their lethal dietary toxicity to young bob-
whites, coturnix quail, ring-necked pheasants,
and mallards (Heath et al. 1972a). Aldrin and
dieldrin were among the six most toxic chemi-
cals of those tested on all species, and toxa-
phene was the only other organochlorine that
was particularly toxic to mallards. Major
species differences in vulnerability to various
chemicals such as were demonstrated in this
study must be considered whenever toxicity
of particular chemicals to avian species is un-
known. Further testing made this point in-
creasingly clear (Hill et al. 1975). Among the
more toxic organochlorine compounds, nearly
all are alicyclic hydrocarbons. Of these chemi-
cals tested, most of the aromatic chlorinated
hydrocarbons are among the less toxic.

Toxicities of six PCB compounds (Aroclor
1232, 1242, 1248, 1254, 1260, and 1262) to
penned mallards, pheasants, bobwhite, and
coturnix quail were generally less than that of
DDT (Heath et al. 1972b). Aroclor toxicity
was positively correlated with chlorine per-
centage (last two digits of Aroclor number) for
the 2-week-old birds that were fed treated
diets for 5 days. The joint toxicity of Aroclor
1254 and DDE on coturnix was additive, not
synergistic. When 18 chemicals (including 8
organochlorines) were fed in 13 pairs to co-
turnix quail and ring-necked pheasant, the
effects of the organochlorines also were addi-
tive rather than synergistic (Kreitzer and
Spann 1973).

To learn if the percentage of chlorine in a

mixture of PCB's alone determines toxicity,
Hill et al. (1974) fed coturnix quail diets con-
taining Aroclor 1248, 1254, or 1260 at levels
that added equal amounts of chlorine to the
feed. Sublethal concentrations produced no
detectable effects. Lethal concentrations with
equal chlorine showed Aroclor 1248 to be the
least toxic of the three compounds at the
highest chlorine concentrations. At lower con-
centrations, Aroclor 1254 was the most toxic
Aroclor. Although chlorine percentage of a
PCB is positively correlated with its avian
toxicity, PCB toxicity is apparently not
simply a function of chlorination. Toxicity
also is related to the positions the chlorine
atoms occupy on the benzene rings. Toxicity
of hexachlorobiphenyl mixtures to bird em-
bryos has been shown to be correlated with
their dibenzofuran content (Vos and Koeman
1970; Vosetal. 1970).

Experiments with coturnix quail were used
to simulate the stresses on wild birds of breed-
ing condition and of weight loss due to migra-
tion (Gish and Chura 1970). Light conditions
in the laboratory were manipulated to stimu-
late reproductive development in one group of
birds and suppress development in another
group. Within each of these groups, some
birds were partially starved before dosage
and some were fully fed. Birds were then fed
dietary levels of 0, 700, 922, 1,214, or
1,600 ppm (dry weight) of DDT for a period of
20 days or until death. Birds partially starved
before dosage were more susceptible to DDT
intoxication than nonstarved ones. Similarly,
males died earlier than females, and the
lighter birds died earlier than the heavier
ones. The heavier birds of each sex not only
survived longer than lighter individuals re-
ceiving the same treatments, but they also
lost a greater proportion of their weight
before death. During the early portion of the
dosage period, females in breeding condition
were less sensitive to DDT than were non-
breeding females and males. After 10 days on
dosage, however, the cumulative mortality of
females in breeding condition rapidly ap-
proached that of males and of females not in
breeding condition.

Reproduction

Field and experimental evidence indicates
that declines in eggshell thickness observed in

certain species in North America and Great
Britain since the mid- 1 940 's have been largely
caused by residues of p,p '-DDE or other com-
pounds or metabolites of the DDT group
(Cooke 1973). At moderate or high levels of
DDE, shell thinning is severe and eggs may
break during incubation. High DDE levels
have been recorded in California; species af-
fected there have included brown pelicans
(Risebrough et al. 1971), double-crested cor-
morants (Gress et al. 1973), great egrets, and
great blue herons (Faber et al. 1972). As indi-
cated previously, much of the DDE probably
originated from an insecticide manufacturing
plant in southern California. DDE levels asso-
ciated with the shell thinning of eggs of the
common murres (Gress et al. 1971) and ashy
storm petrels (Coulter and Risebrough 1973)
on the Farallon Islands in central California
may also have originated in part from this
particular source.

Eggshell thinning has occurred in several
other species that occur in freshwater or es-
tuarine habitats or that nest on coastal
islands. In 1967, shell thickness in herring
gull eggs from five States decreased with in-
creases in chlorinated hydrocarbon residues
(Hickey and Anderson 1968). Comparison of
eggshells taken before 1946 with those taken
since then reveals that several species includ-
ing the peregrine falcon, brown pelican,
double-crested cormorant, black-crowned night
heron, bald eagle, and osprey have sustained
shell-thickness and shell-weight decreases of
20% or more, at least for brief periods (Ander-
son and Hickey 1972). In some of these, re-
gional population declines are known. How-
ever, in seabird species that depend upon ma-
rine food chains in Iceland, there was no evi-
dence of shell thinning in 1973 (J. A. Sproul et
al., unpublished manuscript).

Shell thickness was significantly and in-
versely correlated with the concentration of
DDE in 40 great blue heron eggs from Alberta
(Vermeer and Reynolds 1970; Vermeer and
Risebrough 1972).

In the Upper Great Lakes States, 9 of
13 species of fish-eating birds were found in
1969-70 to have sustained statistically signifi-
cant decreases in eggshell thickness since
1946 (Faber and Hickey 1973). Maximum
changes in a thickness index occurred in great
blue herons (-25%), red-breasted mergansers
(Mergus serrator; -15%), and double-crested

cormorants (-15%). Heron eggs taken in Lou-
isiana generally displayed a smaller post- 1946
change than herons in the Middle West. Al-
though DDE was a prominent factor for most
groups, especially herons, in relation to the
eggshell thinning observed, dieldrin and
PCB's also were associated with thinning in
some species. This relationship, however, may
have been due to correlation in concentrations
of these chemicals and concentrations of
DDE.

The thinning of eggshells of the brown peli-
can has proven to be related to the concentra-
tions of DDE in the eggs (Blus et al. 1971;
Blus et al. 1972a, 1972b). Nearly all brown
pelican eggs collected from 13 colonies in
South Carolina, Florida, and California in
1969 and from 17 colonies in South Carolina
and Florida in 1970 exhibited eggshell
thinning (Blus 1970; Blus et al. 1974a). Of the
100 eggs analyzed for residues of pollutants,
all eggs contained measurable quantities of
DDE; most eggs contained measurable quan-
tities of DDD, DDT, dieldrin, or PCB's. DDE
appears to have been responsible for virtually
all the eggshell thinning.

Nest success of brown pelicans in South
Carolina was related to residues of DDE and
dieldrin in sample eggs (Blus et al. 1974b).
Residues of DDE seemed primarily respon-
sible for nest failure; however, deleterious ef-
fects of this pollutant on nest success was not
satisfactorily separated from those induced
by dieldrin. Significant intercorrelation of all
five organochlorine residues identified in the
eggs complicated the relationship of residues
to nest success. Maximum DDE residues in
an egg from a successful nest were 2.4 ppm
and in an egg from an unsuccessful nest,
8.5 ppm. Comparable maximum residues for
dieldrin in sample eggs were 0.54 ppm (suc-
cessful) and 0.99 ppm (unsuccessful). Resi-
dues of DDD, DDT, or PCB's in sample eggs
were not significantly related to nest success.
Reproductive success in the brown pelican
colony was subnormal in the 2 years of study
(1971 and 1972) but reproductive success was
normal in those nests in which the sample egg
contained either 2.5 ppm or less of DDE, or
0.54 ppm or less of dieldrin.

Residues of DDE, DDD, DDT, dieldrin, and
PCB's exhibited a significant decline in South
Carolina brown pelican eggs from 1969
through 1973 (Blus et al. 1977a), but the de-

crease in DDD was greatest. In 1973, the peli-
cans experienced excellent reproductive
success for the first time in many years, and
the decline in residues was related to this im-
provement. DDE was implicated as the agent
responsible for most pollutant-induced nest
failure; residues above 3.7 ppm in the sample
egg were associated with total failure of those
eggs remaining in the nest. The improvement
in reproductive success was not associated
with an increase in average eggshell thick-
ness.

The peregrine falcon appears to be affected
by shell thinning in all areas of its nearly glo-
bal range thus far examined, including areas
in the Aleutians (Peakall et al. 1975), Green-
land (Walker et al. 1973), and coastal Chile
(Walker et al. 1973) where they depend on ma-
rine food chains. On the Auckland Island in
the sub- Antarctic, one egg of the New Zealand
falcon contained DDE residues that were
similar to those associated with shell thinning
in the closely related peregrine (Bennington et
al. 1975).

Peregrine falcons that breed along the coast
of Scotland feed largely on seabirds, and these
populations have declined in numbers at a
time when populations that were preying on
land birds in the interior remained stable (Rat-
cliffe 1972). A decline in reproductive success
of the white-tailed eagle {Haliaeetus albicilla)
in Germany has most likely been caused by
DDE (Koeman et al. 1972b). In the Baltic,
where white-tailed eagle populations declined
during this century (Henriksson et al. 1966),
very high concentrations of PCB and DDT
compounds have been measured in eagles that
were found dead (Jensen et al. 1972).

During an early study, the population of the
Bermuda petrels (Pterodroma cahow) was
undergoing an unexplained decline that was
attributed to the presence of DDT (Wurster
and Wingate 1968), but reproductive success
subsequently improved. Reexamination of the
tissues that had been analyzed for DDT, and
analysis of dead chicks and unhatched eggs
obtained subsequently, showed no changes in
either DDT or PCB concentrations during the
periods of poor reproductive success and sub-
sequent recovery. Moreover, residues were
comparatively low when related to those of
other species of petrels in more contaminated
areas (D. Wingate and R. W. Risebrough, un-
published data).

Shell thinning of eggs of the osprey in the
northeastern United States where reproduc-
tion has been low and where population
numbers have declined is also related to DDE
concentrations (Spitzer et al. 1977). Dieldrin
and PCB's also may have contributed to the
rapid population decline in the affected areas
in the Northeast, principally Connecticut
(Wiemeyer et al. 1975).

In the Northeast, shell thinning has been
documented in eggs of the gannets breeding
on Bonaventure Island (J. A. Keith, personal
communication). The breeding population of
gannets, after increasing over the previous
80 years, declined by 16% between 1969 and
1973 (Nettleship 1975). In the recent past,
DDT was extensively used in forest spray op-
erations in adjacent areas of New Brunswick.

Patterns of reproductive failure in declining
populations of several European and North
American raptorial species were duplicated
experimentally with captive American kes-
trels that were given a diet containing dieldrin
and DDT, two commonly used organochlorine
insecticides (Porter and Wiemeyer 1969).
Major effects on reproduction were increased
egg disappearance, increased egg destruction
by parent birds, and reduced eggshell
thickness.

In other experimental studies, DDE has
caused significant eggshell thinning in cap-
tive screech owls (Otus asio) (McLane and
Hall 1972) and American kestrels (Wiemeyer
and Porter 1970). The levels of DDE found in
the kestrel eggs in the second reproductive
season of that study are similar to those
found in British peregrine falcon eggs (Rat-
cliffe 1967).

Bald eagle eggs collected in 1968 from nests
in Wisconsin, Maine, and Florida all con-
tained residues of DDE, DDD, dieldrin, hep-
tachlor epoxide, and PCB's (Krantz et al.
1970). Many also contained traces of DDT.
Eggs from five nonproductive nests in Maine
contained much higher residues than did eggs
collected from either productive or nonpro-
ductive nests in Wisconsin and Florida.

Twenty-three bald eagle eggs collected in
Alaska, Maine, Michigan, Minnesota, and
Florida during 1969 and 1970 were analyzed
for organochlorines and mercury (Wiemeyer
et al. 1972). All eggs contained residues of
DDE, dieldrin, PCB's, and mercury. Average
residue concentrations were lowest in eggs

from Alaska. Significant eggshell thinning
has occurred among eggs in samples from
most major areas. Some eggs contained DDE
residues of the same magnitude as those that
produced shell thinning in experimental
species. High dieldrin residues in some eggs
could have an adverse effect on reproductive
success.

Egg failure was the major cause of poor re-
productive success of ospreys on the Potomac
River during 1970 (Wiemeyer 1971). Many
eggs disappeared between visits to the nests;
some were found broken or damaged in the
nests, and others failed to hatch.

Osprey eggs were exchanged between Con-
necticut and Maryland nests in 1968 and 1969
to determine which environmental factors
might have contributed to the decline in re-
productive success of Connecticut ospreys
(Wiemeyer et al. 1975). Incubation of 30 Con-
necticut osprey eggs by Maryland ospreys did
not improve the hatching rate. Forty-five
Maryland osprey eggs incubated by Connecti-
cut ospreys hatched at their normal rate. The
results of the egg exchanges and associated
observations indicated that the most prob-
able cause of the poor reproduction of Connec-
ticut ospreys was related to contamination of
the birds and their eggs. Residues of DDT and
its metabolites, dieldrin, and PCB's were gen-
erally higher in fish from Connecticut than
from Maryland. There were no major changes
in residue content of Connecticut eggs col-
lected in 1968-69 compared with those col-
lected in 1964. One Connecticut osprey had a
concentration of dieldrin in its brain that was
in the lethal range. The average shell thick-
ness of recently collected osprey eggs from
Connecticut had declined 18%, and those
from Maryland had declined 10% from pre-
1947 norms. Dieldrin, DDE, and PCB's are
three environmental pollutants that have
most likely been important factors in the
greatly reduced reproductive success and
rapid population decline of Connecticut
ospreys.

All black duck eggs that were collected in
1971 from the northeastern United States and
Canada contained DDE residues (Longcore
and Mulhern 1973). Means for States and
Provinces ranged from 0.09 to 5.94 ppm, with
mean concentrations exceeding 1.0 ppm in
eggs from Maine, New York, New Jersey, and
Delaware. The highest DDE concentration
(14.0 ppm) was in an egg from Delaware. The

DDD and DDT residues averaged < 0.5 ppm
for each collection area. No mirex residues
and only trace amounts of dieldrin and hepta-
chlor epoxide were detected. Of the 61 eggs,
57 contained PCB's; means ranged from
< 0.05 ppm in samples from Nova Scotia to
3.30 ppm in those from Massachusetts, with
trace amounts occurring in nearly half the
samples. Mean organochlorine pesticide resi-
dues were lower in the 1971 samples than in
those analyzed in an earlier study in 1964.
Average shell thickness of eggs collected in
1964 (0.321 mm) was significantly less
{P < 0.01) than that of eggs collected before
1940 (0.348 mm) or in 1971 (0.343 mm).

Eggs of captive black ducks fed diets con-
taining DDE at 10 and 30 ppm (dry weight;
approximately 3 and 9 ppm wet weight) ex-
perienced significant shell thinning and an in-
crease in shell cracking when compared with
eggs of untreated black ducks (Longcore et al.
1971b). Survival of ducklings from dosed
parents in terms of "percentage of 21-
day ducklings of embryonated eggs" was 40-
76% lower than survival of ducklings from un-
dosed parents. Average DDE residues in eggs
from hens fed 10 and 30 ppm DDE were
46 ppm and 144 ppm.

In another experiment, black duck hens fed
10 ppm (dry weight) of DDE in the diet laid
eggs with shells 22% thinner at the equator,
30% thinner at the cap, and 33% thinner at
the apex than those of controls (Longcore and
Samson 1973). Natural incubation increased
shell cracking more than fourfold as compared
with mechanical incubation. Hens removed
cracked eggs from nests, and one hen termi-
nated incubation. Hens fed DDE produced
one-fifth as many ducklings as did the con-
trols. The DDE in eggs of dosed hens aver-
aged 64.9 ppm.

Concentrations of 10 and 40 ppm DDE (dry
weight) in the feed of penned mallard ducks
caused significant eggshell thinning and
cracking and a marked increase in embryo
mortality (Heath et al. 1969). In other studies,
eggshell thinning also occurred in mallards
fed DDE (Haegele and Hudson 1974), DDT
(Tucker and Haegele 1970; Davison and Sell
1974), or dieldrin (Lehner and Egbert 1969;
MuUer and Lockman 1972; Davison and Sell
1974), but low dietary levels (25 and 50 ppm)
of Aroclor 1254 produced no measurable re-
productive effects (Heath et al. 1972b).

Diets containing various levels of DDT (at
20 ppm, dry weight, or greater), or dieldrin (at
10 ppm, dry weight) caused significant reduc-
tion in eggshell thickness, weight, and cal-
cium in mallard ducks (Davison and Sell
1974). The reduction in eggshell thickness was
linear with increasing dose of dieldrin through
all levels studied.

Mallards were fed untreated feed or feed
containing 40 ppm (dry weight) DDE, 40 ppm
PCB, or 40 ppm DDE + PCB beginning a
month before laying (Risebrough and Ander-
son 1975). Mean shell thickness indices were
similar in the control and PCB groups, but
they were reduced by 17% in the DDE group
and 19% in the DDE + PCB group. The con-
tents of 12 eggs randomly selected from the
DDE group contained 373 ppm DDE (lipid
basis), and 13 eggs from the DDE + PCB
group contained mean residues of 344 ppm
DDE + 364 ppm PCB (lipid basis). Egg pro-
duction was similar in all groups for about the
first 7 weeks, then it dropped markedly in the
DDE + PCB group. Part, but not all, of this
group's lower production of intact eggs was
caused by egg eating. This behavior ac-
counted for 18 of 282 eggs observed lost in
the DDE + PCB group, 6 of 394 eggs in the
PCB group, and none in the control and DDE
groups. Although there was no significant
change in shell thinning or DDE residues
when PCB was added to the diet, the reduc-
tion in the number of intact eggs produced by
the DDE + PCB group suggests that the two
compounds may nevertheless interact to in-
fluence reproductive success.

Behavior

In England, gray herons {Ardea cinerea)
have been observed breaking their own eggs,
and others dropped their live young from the
nest (Milstein et al. 1970; Prestt 1970). Such
aberrant behavior may be related to sublethal
organochlorine residues in the birds, as these
authors suggested. The birds did not eat the
eggshells, but tossed even the fragments from
the nest. Therefore, the alternative possibility
of calcium "hunger" does not seem to be true
in herons.

Mallard ducks fed a diet containing 3 ppm
DDE (dry weight; equal to about 0.6 ppm in a
natural succulent diet) laid eggs that con-
tained an average of 5.8 ppm DDE; duckUngs

that hatched from these eggs differed from
controls in behavioral tests designed to
measure responses to a maternal call and to a
frightening stimulus (Heinz 1976b). In re-
sponse to the maternal call, ducklings from
parents fed DDE were hyper-responsive; com-
pared with controls, a greater percentage ap-
proached the call and a greater percentage of
those that approached remained near" the call
for the remainder of the test. In a test of
avoidance behavior, ducklings whose parents
were fed DDE traveled shorter distances from
the frightening stimulus than did controls.

Coturnix quail chicks were given sublethal
amounts of chlordane, dieldrin, endrin, DDE,
or Aroclor 1254 in their feed, beginning at
7 days of age, and their avoidance response to
a moving silhouette was measured daily for
14 days (Kreitzer and Heinz 1974). The birds
were on dosage for 8 days, and on untreated
feed for 6 days immediately thereafter. Group
avoidance response was significantly sup-
pressed {P from 0.01 to < 0.001) by chlordane,
dieldrin, endrin, and Aroclor 1254, but no
effect of DDE on the birds' behavior could be
detected. The behavior of the endrin-treated
birds returned to normal after 2 days on un-
treated feed. The data indicated partial recov-
ery for birds treated with dieldrin and chlor-
dane, but none for those treated with Aroclor
1254.

Heavy Metals

The sources, occurrence, food web transfer,
and toxicology of heavy metals and other
trace elements must be understood to eval-
uate the significance of these chemicals to ma-
rine birds. These more general aspects have
received considerable attention in recent sym-
posia and reviews (Larsson 1970; Nelson et al.
1971; Gavis and Ferguson 1972; Eisler 1973;
National Research Council of Canada 1974;
Leland et al. 1975). Consequently, our discus-
sion will be restricted to the more specific
aspects of the exposure of aquatic birds to
these chemicals, but will include some inter-
pretive information relative to terrestrial
avian species.

Most techniques that are used for measur-
ing mercury residues in environmental
samples determine levels of total mercury, re-
gardless of the chemical form in which it

occurs. The various forms of mercury, how-
ever, differ widely in their toxicities. Unless
otherwise specified, mercury concentrations
presented here represent concentrations of
total mercury.

Exposure of Marine Birds to
Heavy Metals

Animals acquire heavy metals from the
foods they eat, from the water that surrounds
them, and possibly from the air they breathe.
Quantities accumulated differ greatly among
organisms, depending upon exposure and
physiology (White and Stickel 1975).

Mercury in tissues of living organisms is
often primarily in the more toxic methyl mer-
cury form (Westoo 1967; Fimreite 1974), and
methyl mercury is readily incorporated into
the bodies of aquatic organisms (Leland et al.
1975). Most of the mercury in fish is in the
form of methyl mercury (Koeman et al. 1975),
but the high mercury concentrations dis-
covered in the livers of six dead great cormor-
ants and in hvers of three others that were col-
lected in the Netherlands were not primarily
methyl mercury (Koeman et al. 1973). Mer-
cury concentrations, primarily in forms other
than methyl mercury, increased with age in
some marine mammals and were correlated
with concentrations of selenium and bromine
(Koeman et al. 1975; Martin et al. 1976). Per-
haps, like some marine mammals, cormorants
may be able to detoxify methyl mercury by a
chemical mechanism in which selenium and
bromine are involved. However, mercury and
selenium concentrations in livers of common
murres and of a razorbill (Koeman et al. 1975)
and in liver and breast muscle of sooty terns
of known age were not correlated (P. G.
Connors et al., unpublished manuscript). Inor-
ganic and organic mercury from industrial
sources may be converted into methyl mer-
cury by some organisms, including birds
(Jensen and Jernelov 1969; Kiwimae et al.
1969).

Mercury concentration increases with body
weight, or age, in fish (Bache et al. 1971; Fim-
reite et al. 1971), crayfish (Vermeer 1972), and
herons (Hoffman 1974). The concentration in-
creases at higher trophic levels in fish, other
aquatic organisms, fish-eating birds, or ducks
(de Goeij 1971; Fimreite et al. 1971; Fimreite

1974; Hoffman 1974; Kleinert and DeGurse
1972; Vermeer et al. 1973; Baskett 1975).

Mercury concentrations in various tissues
of the body are correlated with each other
(Fimreite 1971; Koeman et al. 1971; Vermeer
and Armstrong 1972a; Fimreite 1974; Heinz
1974, 1976a; Hoffman 1974). Eggs normally
contain between a fifth and a ninth of the mer-
cury concentration in the liver of the female
(Fimreite et al. 1970; M. T. Finley, personal
communication). Mercury in the liver of fe-
male California gulls (Larus californicus) aver-
aged 5.5 times that in their eggs (Vermeer
1971a).

High mercury residues in aquatic
organisms and in the related avifauna are
often related to discharges from chlor-alkali
plants, pulp mills, or other industrial plants
that use mercury (Fimreite 1970; Fimreite et
al. 1971; Nelson et al. 1971; Vermeer 1971a).
Ospreys and great crested grebes (Podiceps
cristatus) now have about 3 times as much
mercury in some industrially contaminated
areas as in uncontaminated areas (Larsson
1970).

In a survey of aquatic birds at 33 locations
in Alberta, Saskatchewan, and Manitoba,
mercury levels were generally higher in gulls
{Larus spp.) and fish-eating birds than in
ducks and geese (Vermeer 1971a). The highest
mercury levels were found in herring gulls,
possibly related to their scavenging and fish-
eating habits.

Elevated mercury levels were found in
livers of common mergansers {Mergus mer-
ganser; up to 86 ppm), common loons
(90ppm), and great blue herons (128 ppm)
from Ontario (Fimreite 1974). Lower concen-
trations were found in mallards (12.5 ppm)
and pintails (Anas acuta; 6.2 ppm). Mercury
levels were higher in adults than in imma-
tures. A chlorine plant about 80 km upstream
from the collecting locahty was believed to be
the source of mercury found in the birds.

Mercury was present in spotted sandpiper
{Actitis macularia) eggs collected upstream
from Edmonton, Alberta, at lower concentra-
tions (0.09 ppm) than in those eggs collected
downstream (0.28 ppm), suggesting municipal
or industrial contamination originating at
Edmonton (Vermeer 1971b).

During another survey in Canada, highest
concentrations of mercury in livers of fish-eat-
ing birds collected near sites of industrial con-
tamination were in red-necked grebes

(Podiceps grisegena; Fimreite et al. 1971).
Four common tern eggs averaged 0.58 ppm
and two red-breasted merganser eggs aver-
aged 0.81 ppm.

Aquatic bird eggs from the upper Great
Lakes States contained higher mercury levels
than those from Louisiana, although species
represented from the two areas were not iden-
tical (Faber and Hickey 1973). Highest mean
residues were in three species of mergansers
(up to 1.6 ppm; red-breasted merganser). For
those species with all eggs containing less
than 0.25 ppm of mercury, the residues were
considered to represent background levels.
Mercury exceeded 1 ppm in one or more eggs
of black-crowned night heron, hooded mer-
ganser (Lophodytes cucullatus), common mer-
ganser, and red-breasted merganser. Highest
levels (up to 1.9 ppm) were in addled eggs of
red-breasted mergansers.

Many birds dependent upon aquatic areas
in the Lake St. Clair, Michigan, region have
high residues of mercury in their tissues
(Dustman et al. 1972). In 1970, carcasses,
livers, and eggs were collected and analyzed.
Mercury levels in great blue herons (up to
175 ppm in the liver; 23 ppm in the carcass)
and common terns (up to 39 ppm in the liver;
7.5 ppm in the carcass) far exceeded those in
any other species. The levels are comparable
to those in birds in Sweden that died under ex-
perimental dosage with methyl mercury and
in birds that died under field conditions in sev-
eral Scandinavian countries with signs of mer-
cury poisoning (Henriksson et al. 1966; Borg
et al. 1969; Holt 1969). Mercury residues in
eggs of all of the five common terns (up to
6.2 ppm), five of nine mallards (up to 2.7 ppm),
three of the five black-crowned night herons
(up to 1.1 ppm), and the single egg of a pied-
billed grebe (Podilymbus podiceps; 4.0 ppm)
were in the range of residues (0.5-3.1 ppm) in
eggs of ring-necked pheasants whose repro-
ductivity was reduced by mercury in experi-
mental studies (Borg. et al. 1969; Fimreite
1971; Spannetal. 1972).

In 1973, eggs of some of these species were
again collected at Lake St. Clair, following re-
strictions on industrial discharges of mercury
into the St. Clair River (Stendell et al. 1976).
Mercury levels in the eggs were appreciably
lower than were found in these species in
1970. Common terns contained the highest
residues (up to 1.3 ppm). Mallard eggs con-

tained relatively low residue levels (<0.05 to
0.26 ppm). Black-crowned night heron eggs
(up to 0.76 ppm) and great egret eggs (up to
0.45 ppm) contained intermediate amounts.

Mercury levels generally are low in most
species of ducks and geese but higher levels
have been found in those species that con-
sume a greater proportion of animal material
in their diet (Kleinert and DeGurse 1972;
Krapu et al. 1973; Fimreite 1974; Heath and
Hill 1974). Among North American waterfowl
species, the highest levels have been found in
mergansers. Common mergansers from On-
terio had up to 86 ppm mercury in their livers
(Fimreite 1974). Hooded mergansers from
Clay Lake, Ontario, contained up to 12.3 ppm
and common goldeneyes (Bucephala clangula)
up to 7.8 ppm in their breast muscle (Vermeer
et al. 1973). Food items were also analyzed
and crayfish {Oronectes virilis), which the
hooded mergansers eat, contained the highest
average concentration of mercury (7.1 ppm).

Mercury has been found in the visceral fat
of black-footed albatrosses and Laysan alba-
trosses from Midway Atoll, North Pacific
Ocean (Fisher 1973). Average residue levels in
the Laysan albatrosses were 0.104 ppm and
those in the black-footed were 0.075 ppm.

Mercury has been found in the hvers of
birds collected around the British coast (Dale
et al. 1973). The highest concentration
(26 ppm; converted from 122 ppm dry weight,
see Holdgate 1971) was in a red-breasted mer-
ganser. Common eiders, which feed on
mussels that are known to accumulate mer-
cury, also had high concentrations (10 ppm).
All of the more pelagic species, including
black-legged kittiwakes, fulmars, and auks
{Alca torda and Alle alle) had less than
2.2 ppm. Three gannets had slightly higher
levels (up to 2.9 ppm). Herring gulls from
oceanic islands contained relatively low mer-
cury residues (up to 2.6 ppm) like the pelagic
birds, but those from near shore had higher
residues.

Common puffins collected around the coast
of Britain contained up to 7.7 ppm (Parslow et
al. 1972), and eiders from the Tay region had
up to 0.45 ppm mercury in their livers (Jones
etal. 1972).

Elevated levels of mercury have been found
in birds of the Baltic region (Jensen et al.
1972). In 1969, mercury content of common
murre secondaries had doubled the levels
from 1906-1325. Mercury levels in murre eggs

were approximately the same in 1968 and
1969, averaging about 0.52 ppm, with the
upper extreme concentration of 0.67 ppm.
Even higher levels of mercury (3.7 ppm) were
found in muscle tissue of great cormorants
than in muscle of murres (0.9 ppm) or black
guillemots (1.8 ppm) from the Baltic. Mercury
in feathers (up to 51 ppm), muscle (up to
26 ppm), and brains (up to 14 ppm) of white-
tailed eagles exceeded the levels in other
species. Mercury concentrations in the kid-
neys (48-123 ppm) and muscle tissue (1.9-

8.5 ppm) of other white-tailed eagles from the
same area further indicate that the species
may have serious mercury pollution problems
(Henriksson et al. 1966). Bald eagles in the
United States also occasionally contain high
levels of mercury (up to 43 ppm) in their car-
casses (Belisle et al. 1972).

Mercury levels (figures not specifically
stated) in the muscle of eiders and sandwich
terns of the Dutch Wadden Sea appear 3 to
5 times higher than the levels considered rep-
resentative of natural background (de Goeij
1971). Analyses were also made of various
organs of three common murres and one
razorbill that were found as oiled birds along
the Dutch coast (Koeman et al. 1975): mer-
cury in the livers did not exceed 2.5 ppm; sele-
nium in the liver of one common murre was

4.6 ppm, but the levels of these metals were
not correlated with each other.

There were no significant geographical or
species differences in two essential heavy
metals (copper and zinc) in Antarctic and
North American petrels (Anderlini et al.
1972). Silver, cobalt, and lead were difficult to
detect at the low levels that were found, but
there were no detectable differences in their
concentrations. Cadmium, chromium, nickel,
and mercury levels in petrels suggested a cor-
relation of increasing concentration with in-
creased exposure to industrialized areas.
Higher concentrations of these metals in ashy
petrels are probably the result of their feeding
in the proximity of San Francisco Bay.

Livers from ruddy ducks killed by an oil
spill on the Delaware River contained detect-
able levels of lead, cadmium, and mercury
(White and Kaiser 1976). Lead ranged from
0.19 to 0.61 ppm, cadmium from 0.27 to
1.60 ppm, and mercury from 0.06 to
0.74 ppm. Residues of these metals were simi-
lar to those found in canvasbacks from the

Chesapeake Bay region (D. H. White and
R. C. Stendell, unpublished manuscript).

Mercury residues in the livers of six gannets
from the Irish Sea (4 ppm; 18.4 ppm dry
weight) averaged higher than in two gannets
from eastern Scotland (1.6 ppm; 7.3 ppm dry
weight) that died during unrelated large-scale
mortality incidents (Parslow et al. 1973).
Average levels of copper (7.4 ppm; 34 ppm
dry weight) and zinc (64.8 ppm; 298 ppm dry
weight) in the livers of gannets from the Irish
Sea also were higher than in two others from
eastern Scotland (2.8 ppm copper; 26.3 ppm
zinc). The differences in the metal concentra-
tions between the two groups were considered
the result of the differences in liver sizes. Al-
though the cause of the gannet deaths could
not be established, heavy metal concentra-
tions in the birds apparently were responsible
for the death of only one individual with high
mercury levels (22 ppm; 98 ppm dry weight).
Lead and cadmium concentrations were below
the limits of detection in all of these birds, but
another gannet that died in an earlier incident
had measurable residues of lead (0.2 ppm) and
cadmium (2.0 ppm).

Mercury concentrations in the livers and
kidneys of common murres that died in the
seabird wreck in the Irish Sea during autumn
1969 did not exceed 5 ppm (23 ppm dry
weight) (Holdgate 1971). Some of the birds
showed relatively high levels of particular
metals and in some the highest concentra-
tions were above the level at which poisoning
may have occurred. However, in general, the
range of mercury levels in the casualties of the
incident and in the healthy birds shot for com-
parison overlap. The levels of mercury (up to
5 ppm), lead (8.7 ppm), cadmium (2.8 ppm),
and arsenic (8.3 ppm) in the livers and kid-
neys of some birds appeared elevated.

Biological Effects of Heavy
Metals on Marine Birds

Toxicology, Physiology, and Pathology

In 1953 a severe neurological disorder
caused by mercury poisoning was first recog-
nized among people living in the vicinity of
Minamata Bay, Japan (Kurland et al. 1960).
Toxic effects and similar histopathological
changes have been reported for fish, birds,
and mammals that died as a result of mercury

poisoning, but no particular studies have been
made on the toxicity of heavy metals to sea-
birds (Parslow et al. 1973). However, in
certain terrestrial species, symptoms of poi-
soning might be expected when mercury con-
centrations in liver or kidney tissues reach
about 30ppm (W. H. Stickel 1971). By con-
trast, normal levels are less than 1 ppm.

Although death may not have been caused
by mercury poisoning, mercury in the livers of
adult great egrets found dead in California
ranged between 2 and 9.5 ppm (Faber et al.
1972). Mercury in the liver (22 ppm; 98 ppm
dry weight) of a gannet from the Irish Sea
could have caused the bird's death (Parslow et
al. 1973).

Female mallards fed 3 ppm mercury (dry
weight) as methyl mercury in their diet had
average mercury residues of 11.1 ppm in their
livers, 14.7 ppm in their kidneys, 5.0 ppm in
their muscles, 4.6 ppm in their brains, and
5.5 to 7.4 ppm in their eggs (Heinz, 1976a).
Males had higher residues, and many of the
duckhngs from these parents died within
1 week after hatching (Heinz 1974, 1976a).
The ducklings also had high levels of mercury
in their tissues.

In short-term tests of lethal dietary toxicity
of pesticidal chemicals, Ceresan M, a fungi-
cide containing ethyl mercury, was relatively
more toxic to young mallards than were 37
other compounds (Heath et al. 1972a). Only
endrin and Dasanit were more toxic. In simi-
lar subsequent tests, Morsodren, another
organomercurial fungicide, was also highly
toxic to young mallards (Hill et al. 1975).

Mercury potentiated the toxicity and
biochemical effects of parathion in coturnix
quail fed a sublethal concentration of Morso-
dren (4 ppm dry weight as methyl mercury)
for 18 weeks (Dieter and Ludke 1975). Mean
residue concentrations in these birds were
21 ppm of mercury in the liver and 8.4 ppm in
the carcass. The computed LD50 of parathion
was 5.86 mg/kg in birds not fed Morsodren
and 4.24 in those fed the heavy metal. When
challenged with a sublethal oral dose of para-
thion (1.0 mg/kg), Morsodren-fed birds ex-
hibited significantly greater inhibition of
plasma and brain cholinesterase activity than
controls.

After administration of various mercury
compounds to domestic chickens (Gallus
gallus), the methyl mercury compounds were

rather evenly distributed among the organs,
whereas the other mercury compounds, or-
ganic and inorganic, gave very high concen-
trations in the liver and kidneys compared
with other organs (Kiwimae et al. 1969). Dif-
ferences in the proportion of methyl mercury
compounds to total mercury occurred in the
white and the yolk of the eggs from these
hens. Although the proportion in the white
was similar to that in the blood and the
muscles, the proportion in the yolk was simi-
lar to that in the liver and the kidneys. The al-
bumen contained mainly methyl mercury
compounds in concentrations that varied with
the compound given to the hens. The methyl
mercury concentration in albumen was
always much lower when other compounds
were administered than when the hens were
given the same quantity of methyl mercury
hydroxide.

In another study, mercury was not detect-
able in the albumen but was present at high
levels in the yolk following intravenous injec-
tion of mercuric nitrate into laying coturnix
quail (Nishimura et al. 1971).

Evaluation of enzymatic profiles appears to
be a potentially valuable technique for moni-
toring the presence of toxicants in wild popu-
lations, especially if used to complement
standard chemical residue analysis (Dieter
1975). Lactate dehydrogenase activity in-
creased twofold and cholinesterase activity
decreased in birds fed Morsodren. After feed-
ing for 3 weeks, mercury in starling carcasses
reflected the concentrations fed daily,
whereas the concentration in the livers was
2 to 4 times that in the diet.

A decrease in cholinesterase activity oc-
curred in male coturnix quail that were fed
diets for 12 weeks containing graded levels of
mercuric chloride (Dieter 1974). At 12 weeks
the decrease was proportional to the log dose
received, although this was not true after
2 and 4 weeks on the treated diet. Mercury
residues attained in the tissues were 5% or
less of those in the feed.

There was a marked sexual difference in
rates of mercury loss in coturnix quails (Back-
strom 1969). Males lost little of the mercury
in their bodies, especially from the brain and
muscle, in 30 days, but females had a marked
loss in this period, largely because of excre-
tion in eggs. Ring-necked pheasants lost 33%
to 50% of the mercury from their livers and

kidneys in 2 months, and approximately 99%
was lost in 6 months (Borg et al. 1969).
Ospreys apparently have a similar loss rate of
mercury (Johnels et al. 1968).

When methyl mercury dicyandiamide was
fed to mallard ducks at a concentration of
3 ppm mercury (dry weight), mercury accumu-
lated in the eggs to an average of 7.2 and
5.5 ppm in 2 successive years (Heinz and
Locke 1976). Mercury in the eggs caused brain
lesions in ducklings. Lesions included demye-
lination, neuron degeneration, necrosis, and
hemorrhage in the meninges overlying the
cerebellum. Brains of dead ducklings con-
tained an average of 6.2 and 5.2 ppm mercury
in the 2 successive years.

Upon necropsy, ring-necked pheasants that
were killed after receiving 4.2 ppm mercury
(dry weight) in their diet for 350 days ap-
peared normal, but those that received
greater concentrations (12.5, 37.4, or
112 ppm, dry weight) died during the experi-
ment and showed variable amounts of subcu-
taneous edema and decreasing amounts of
subcutaneous and abdominal adipose tissue
as survival time on the treated diet increased
(Spann et al. 1972), Birds that died on the
higher dosages showed signs of neurological
disturbance, including ataxia and torticollis,
before death.

Lead poisoning has long been recognized as
a serious problem for waterfowl (Wetmore
1919; Jordan and Bellrose 1951; Bellrose
1959). Histopathological changes occur in the
kidneys of mallards as a result of lead shot in-
gestion (Locke et al. 1966, 1967). In addition,
significant changes in activity of three en-
zymes often used to assess hepatic damage
occurred in mallard ducks following oral ad-
ministration of lead shot (Rozman et al. 1974).

The ingestion of one number 4 lead shot by
each of 80 pen-reared mallards that were fed
whole-kernel corn caused 19% mortality
within an average of 20 days (Longcore et al.
1974a). Coating or alloying lead with other
metals only delayed mortality among dosed
ducks. Disintegrable lead shot with water-
soluble binder and lead-containing biochemi-
cal additives were as toxic to mallards as com-
mercial lead shot.

Lead levels in brains, tibiae, and breast
muscle of mallard ducks that died and in
tibiae of those that were sacrificed increased
significantly from dosage with one number 4
lead shot (about 1.4 g) until death (Longcore

et al. 1974b). In mallard ducks, lead levels ex-
ceeding 3 ppm in the brain, 6 to 20 ppm in the
kidney or liver, or 10 ppm in clotted blood
from the heart indicated acute exposure to
lead.

One month after dosage, mean lead levels in
mallards given one number 4 all-lead shot
were about twice those in tissues of mallards
given one number 4 lead-iron shot that con-
tained about 50% lead (Finley et al. 1976a).
Necropsy of sacrificed ducks failed to reveal
any of the tissue lesions usually associated
with lead poisoning in waterfowl. Lead in the
blood of ducks dosed with all-lead shot aver-
aged 0.64 ppm, and 0.28 ppm in ducks given
lead-iron shot. Lead residues in livers and
kidneys of females given all-lead shot were
significantly higher than in males. In both
dosed groups, lead levels in wingbones of the
females were about 10 times those in males,
and were significantly correlated with the
number of eggs laid after dosage. It appeared
that after the laying hens ingested sublethal
amounts of lead shot, high lead deposition in
the bone occurred as a result of mobilization
of calcium from the bone during eggshell for-
mation. Lead levels in contents and shells of
eggs laid by hens dosed with all-lead shot were
about twice those in eggs laid by hens dosed
with lead-iron shot. Lead levels in eggshells
best reflected levels of lead in the blood.

The inverse correlation between delta-
aminolevulinic acid dehydratase (ALAD) ac-
tivity and blood lead concentrations was
highly significant in canvasback ducks from
the Chesapeake Bay (Dieter et al. 1976).
ALAD is an important enzyme in hemoglobin
synthesis. The activity of this enzyme in the
blood provides a sensitive and precise esti-
mate of lead contamination in waterfowl. In
mallards, lead concentrations in blood were
strongly correlated with erythrocyte ALAD
activity, suggesting that biochemical re-
sponse to two types of lead shot (one all-lead,
the other containing 50% lead) depends upon
the quantity of lead present (Finley et al.
1976b).

Reproduction

Mercury levels (3.5 to 1 1 ppm) in the eggs of
Swedish white-tailed eagles that failed to
hatch indicate that the decline in reproduction
of this species could be attributed to mercury
poisoning (Borg et al. 1969). A corresponding

decline in this species in Finland also was as-
sociated with mercury contamination (Hen-
riksson et al. 1966). However, as discussed
earlier, organochlorines also may be partially
responsible for the observed decline.

There were apparently no young produced
by common loons in Clay and Ball Lakes, On-
tario, in 1970 and 1971 (Fimreite 1974). (Both
lakes receive effluent from a chlorine plant.)
Fledging success of common terns at Ball
Lake was 10% of normal, but fledging was
normal at nearby Wabigoon Lake, where
birds contained lower residues. Average total
mercury in the eggs was 3.6 and 1.0 ppm;
average methyl mercury was 2.4 and 0.8 ppm
in the two colonies.

There are considerable differences between
species in susceptibility to mercury pollu-
tants. Mercury concentrations as high as
16 ppm in western Ontario herring gull eggs
apparently did not affect their hatchability
(Vermeer et al. 1973), but 0.5 to 1.5 ppm mer-
cury in ring-necked pheasant eggs reduced
hatchability, reduced egg weight and produc-
tion, and produced a large number of eggs
without shells (Fimreite 1971).

Concentrations of mercury found in the
livers of abnormal young terns (Sterna
hirundo and S. dougallii) from Great Gull
Island (in Long Island Sound) ranged from
0.2 to 1.2 ppm, but were not thought to have
caused the abnormalities (Hays and Rise-
brough 1972). Livers of normal young terns
were not analyzed. Hatchability in the Great
Gull Island colony has consistently been
greater than 90%, but hatchability of com-
mon tern eggs in Lake Ontario colonies has
been low. Concentrations of heavy metals in
common terns were studied to determine the
reason for the difference in hatchability. Con-
centrations of cadmium, chromium, cobalt,
copper, lead, mercury, nickel, silver, and zinc
in bone, liver, breast muscle, and kidneys of
adult birds from the two locations were simi-
lar (Conners et al. 1975). Therefore, these
metals apparently were not responsible for
the differences in hatchability.

Although the reproductive effects of mer-
cury in other species are largely unknown,
mercury residues of 0.5 ppm were associated
with poor reproductive success in an experi-
mental study with ring-necked pheasants
(Fimreite 1971). Average mercury residues in
field-collected eggs of four species of aquatic-
related birds on the Niagara peninsula, On-

tario, were between 0.5 and 1 ppm (Frank et
al. 1975). These included red-winged blackbird
{Agelaius phoeniceus; 0.68 ppm), herring gull
(0.74 ppm), black-crowned night heron
(0.64 ppm), and common tern (0.83 ppm).

Mercury was found in measurable quanti-
ties in all of the 100 brown pelican eggs from
13 colonies in South Carolina, Florida, and
CaUfornia (Blus et al. 1974a). Six of the 21
pelican eggs from South Carohna contained
0.5 ppm or more of mercury. Sixteen of the 49
pelican eggs from Florida contained 0.5 ppm
or more of mercury, and 1 on the verge of
hatching contained 1.43 ppm.

Methyl mercury at low dietary levels (0.5 or
3.0 ppm, dry weight, equal to about 0.1 or
0.6 ppm mercury on the basis of a natural suc-
culent diet) caused lowered reproductive
success in experimental mallards and black
ducks. Mallards fed 3 ppm mercury in their
diet during one reproductive season showed
reproductive impairment, but none was
evident among birds fed 0.5 ppm (Heinz
1974). Adverse effects in the group fed 3 ppm
included a decrease in egg laying, an increase
in embryonic mortality, and reduced duckling
survival. These effects resulted in the produc-
tion of less than half (46.5%) as many 1-week-
old ducklings as the controls. Levels of mer-
cury reached about 1 ppm in eggs of the birds
fed 0.5 ppm mercury and between 6 and
9 ppm in the eggs from ducks fed 3 ppm
mercury.

The hens from the first reproductive season
were kept on diets containing mercury into a
second season (Heinz 1976a). During the sec-
ond season, levels of mercury in eggs from
hens on these diets averaged 0.79 and
5.46 ppm. On a dry-weight basis, the concen-
tration of mercury in eggs was about 5 times
that in the feed. There were no significant dif-
ferences in egg production or hatching suc-
cess among control birds and those fed mer-
cury. However, duckling survival decreased:
ducklings from hens fed 3 ppm mercury
during the two reproductive seasons were less
likely to survive to 1 week of age than were
controls or ducklings from parents fed
0.5 ppm mercury.

Mallards whose parents were fed a diet con-
taining 0.5 ppm mercury (dry weight) were
themselves fed a diet containing 0.5 ppm mer-
cury (dry weight) from 9 days of age through
their first reproductive season (Heinz 1976c).
Mercury in the eggs of these hens fed mercury

averaged 0.86 ppm. Hens fed mercury made
less efficient use of feed and laid a greater per-
centage of their eggs outside their nest boxes
compared with controls. They also produced
fewer 1-week-old ducklings than did controls,
although there had been no difference in duck-
ling production by their parents fed 0.5 ppm
mercury in the preceding years. The ducklings
from dosed parents did not grow as fast as did
those from controls.

Black ducks given a diet containing 3 ppm
mercury (dry weight) as methyl mercury
hatched fewer eggs than did controls and
fewer of their ducklings survived (Finley and
Stendell 1978). Average mercury residues in
brain, Uver, and muscle of ducklings that died
(3.7, 9.4, and 4.9 ppm) were about twice those
in tissues of ducklings sacrificed at 4 weeks of
age (1.6, 5.7, and 2.1 ppm).

Mercury residues in pheasant eggs were 0.9
to 3.1 ppm following administration of
4.2 ppm mercury (dry weight) in their diet
(Spann et al. 1972). The birds exhibited
greatly reduced egg production and increased
embryo mortality in the few eggs laid.

Mercury residues in bobwhite eggs from
birds fed a dietary concentration of 1.7 ppm
(dry weight; administered as ethyl mercury
p-toluene sulfonanilide) averaged 1.6 ppm
(J. W. Spann and R. G. Heath, unpublished
manuscript). There was a significantly greater
mortality among young whose parents re-
ceived mercury in the diet. The principal
period of increased mortality included the last
5 days of incubation and the first day arfter
hatching.

Behavior

The behavior of mallard ducks whose par-
ents were fed a control diet or a diet contain-
ing 0.5 or 3.0 ppm mercury (dry weight) as
methyl mercury was studied (Heinz 1975).
There was no significant difference among
controls and ducklings from mercury-treated
parents in the percentage of ducklings that
approached the tape-recorded maternal call.
However, control ducklings moved back and
forth toward the call more than ducklings
from mercury-treated parents and also spent
more time in the end of the runway near the
loudspeaker than ducklings whose parents
were fed a diet containing 0.5 ppm mercury.

Compared to control ducklings, those from
parents fed a diet containing either mercury
concentration were hyper-responsive in avoid-
ance behavior tests.

Among mallard ducklings produced in the
2nd year of the study in which hens were fed a
control diet or a diet that contained 0.5 or
3 ppm mercury (dry weight), the findings were
similar (Heinz 1976a). There were no signifi-
cant differences among controls and groups
fed mercury in approach responses toward a
recorded maternal call and ducklings from
mercury-treated parents were hyper-respon-
sive compared with controls in avoidance
behavior.

In the second generation, there were no sig-
nificant differences between controls and
ducklings from parents fed 0.5 ppm mercury
in approach responses to tape-recorded
maternal calls, in avoidance of a frightening
stimulus, or in open-field behavior (Heinz
1976c).

Plastic and Other Artifacts

Small plastic beads and irregularly shaped
particles up to 0.5 cm in diameter are com-
monly found in plankton samples from widely
separated oceanic areas, including the north-
western Atlantic, Sargasso Sea, Bristol
Channel (United Kingdom), and the coastal
waters of southern New England (Carpenter
and Smith 1972; Carpenter et al. 1972; Morris
and Hamilton 1974; Colton et al. 1974). The
particles are primarily composed of polysty-
rene or polyethylene compounds and have
about the same density as seawater. Their
various colors include white, green, brown,
blue, red, or clear (Carpenter and Smith 1972;
Morris and Hamilton 1974; Colton et al.
1974). The polystyrene spherules evidently
are of industrial origin, because they have
been found in the effluents from manufacture
of polystyrene (Hayes and Cormons 1974;
Morris and Hamilton 1974). Their abundance
in the British Channel water was lowest near
the seaward end and greatest in the inner part
of the Channel, near the Holm Islands. Ben-
thic sediments near the Holm Islands con-
tained as many as 20,000 beads/m^ (Morris
and Hamilton 1974).

Small fish ingest the beads and particles
(Carpenter et al. 1972; Kartar et al. 1973).

Plastic particles have been found in the
stomachs of fork-tailed petrels, horned
puffins (Fratercula corniculata), and parakeet
auklets (Cyclorrhynchus psittacula) from the
Aleutians (G. J. Divoky and C. M. White, per-
sonal communication), as well as in the
stomachs of adult and nestling Leach's
petrels from Newfoundland and New Bruns-
wick (Rothstein 1973).

Gulls and terns regurgitate indigestible
parts of their food, such as bits of shell and
fish bones. Polystyrene particles have also
been found in these pellets (Hays and Cor-
mons 1974).

It is not known whether the birds ingest the
plastic particles directly, but petrels appar-
ently do. Other marine birds may acquire par-
ticles in their stomachs by consuming fish
that have previously ingested the plastic
particles.

Evidence of harmful effects of plastic par-
ticles to any species is lacking, except for the
possibility of intestinal blockage in smaller
fish (Carpenter et al. 1972). However, they do
accumulate in the environment, are eaten by
fish, and are found in the stomachs of marine
birds. It has been suggested that the plastics
industry develop products that are degrad-
able, but the most likely outcome of such an
effort would be introduction of finished prod-
ucts that would disintegrate into smaller
particles similar to those described here (Hays
and Cormons 1974).

Rubber thread cuttings may represent a
hazard to marine birds. Common puffins, in
particular, appear to mistake them for fish
and swallow them. These elastic threads form
knots and the tangled mass may remain in the
stomach. In one case the entangled elastic
was tightly packed into the gizzard exit; in an-
other it had formed a ball of rubber in the
gizzard itself. Although the rubber threads
may not kill the birds, there is a possibility
that they make them less able to withstand
other stresses (Par slow and Jefferies 1972).

Although other artifacts, such as trash
scattered on beaches or jetsam washed
ashore, may contribute significantly to the
mortality of certain species of marine birds
(Gochfeld 1973), in other circumstances, such
debris may enhance the habitability of an
area. An apparent increase in the number of
black guillemots breeding in the Barrow,
Alaska, area appears to be associated with the

local increase in man-made debris. The birds
tjT)ically nest in cavities in rock cliffs and
crevices in talus slopes. Because such nest
sites are absent in the Barrow area, guille-
mots have nested in an empty oil drum, under
a collapsed building, and under other types of
man-made debris (Divoky et al. 1974).

No explanation has been found for the ap-
pearance along the Northumberland (United
Kingdom) coast of severely debilitated
common murres whose plumage has been ex-
tensively abraded. Fluoride, discharged by a
nearby aluminum smelter, was considered a
possible cause because the birds had a strong
odor resembling chlorine, another closely re-
lated halogen compound. There also were
similarities between the signs observed in the
affected birds and those observed in cases of
acute or chronic fluorosis in other animals.
The implication of fluoride was dismissed,
however, in part, on the basis of low fluoride
residue levels in bone, skin, internal organs,
and digestive tract of the affected birds. Fur-
ther, normal murre feathers were not dam-
aged by soaking in various fluorine com-
pounds, in samples of smelter effluent, and in
undiluted scrubber liquid (Croxall 1972).

Recommendations

The levels of any pollutant, or combination
of pollutants, in the marine environment
should remain below a level that damages the
viability of any population or species of ma-
rine bird. Thus the global use of organo-
chlorine compounds must be regulated, if nec-
essary, to restrict input into the sea. The
undersea exploitation of petroleum, the
marine transport of petroleum, and the ac-
tivities of coastal refining and petrochemical
industries must also be regulated to prevent
harm to local populations of marine birds.

Much remains to be learned about the expo-
sure of marine birds to environmental pollu-
tants in northern North America. The most
critical areas for study include the effects of
chronic sublethal exposure to petroleum hy-
drocarbons, certain organochlorines, and mer-
cury. The possible synergistic effects of these
compounds in marine birds should also be in-
tensively studied.

A long-term program to monitor increasing
or decreasing levels of any particular poUu-

tant in the marine environment, with particu-
lar reference to the levels that affect the most
sensitive species of marine bird, is necessary.
A portion of this program might be carried
out by using the eggs of marine birds, because
colonies of some species are large and eggs
may be obtained on a regular basis. The
variance of pollutant distributions and the
mathematical nature of these distributions
are imperfectly known and the statistics of
sampling have not yet been adequately formu-
lated. Moreover, it would be desirable to carry
out such programs in conjunction with other
programs that examine changes in pollutant
levels in the marine environment Uke the
"Mussel Watch" (Goldberg 1975), which is
following changes in the levels of plutonium
isotopes, petroleum compounds, chlorinated
hydrocarbons, and selected metals in mussels
from U.S. coastal localities.

Priorities in future research might be given
to more intensive studies within local areas to
obtain a better understanding of the
dynamics of pollutant accumulation by birds.
Of primary concern is the need to determine
whether petroleum compounds are accumu-
lated in food webs, including marine birds,
and whether such compounds exert dele-
terious physiological effects. Because pe-
troleum compounds seem to have longer-last-
ing effects in colder water, the impending ex-
ploitation of oil resources in the offshore and
North Slope areas accentuates the urgent
need for information on the environmental
consequences of chronic as well as acute
exposure.

The environmental effects of small plastic
particles that are commonly found in oceanic
areas, including northern North America,
should be investigated.

The relationships between chronic exposure
to environmental pollutants and other envi-
ronmental stresses are relatively unknown, as
are relationships and effects of pollutants on
many of the essential organisms in the food
webs upon which marine and estuarine birds
depend.

An annual symposium on the marine birds
of northern North America should be held to
serve as a forum for presentation of new infor-
mation. The symposium would contribute sig-
nificantly to conservation of the area's nat-
ural resources by facilitating exchange of in-
formation and coordination of further
research.

Acknowledgments

D. J. Snyder assembled much of the re-
quired literature and also typed the manu-
script. M. T. Finley, J. L. Ludke, L. F. Stickel,
and D. H. White reviewed the manuscript and
offered useful suggestions.

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GPO 845 - 756

As the Nation's principal conservation agency, the Department of the
Interior has responsibility for most of our nationally owned public
lands and natural res )urces. This includes fostering the wisest use of
our land and water resources, protecting our fish and wildlife, preserv-
ing the environmental and cultural values of our national parks and
historical places, and providing for the enjoyment of life through out-
door recreation. The Department assesses our energy and mineral
resources and works to assure that their development is in the best
interests of all our people. The Department also has a major rcsiMinsi-
bility for American Indian reservation communities and for people who
live in island territories under U.S. administration.

UNITED STATES

DEPARTMENT OF THE INTERIOR

FISH AND WILDLIFE SERVICE

EDITORIAL OFFICE

AYLESWORTH HALL, C8U

FORT COLLINS. COLORADO 80523

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U.S. DEPARTMENT OF THE INTERIOR

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